the
Add Deposition
RBseaichPragram
BIOPHYSICAL RESEARCH
ACIDIC DEPOSITION AND THE
ENVIRONMENT: A LITERATURE
OVERVIEW
by:
A.H. Legge
R.A. Crowther
Kananaskis Centre for Environmental Research
The University of Calgary
Calgary, Alberta, Canada
November, 1987
PRIME RESEARCH CONTRACTOR: y^^j^j
The Kananaskis Centre for Environmental Research rk«r^rtr.i*i««
The University of Calgary UepOSIIiOn
Calgary, Alberta, Canada Research Program
Digitized by the Internet Archive
in 2015
https://archive.org/details/acidicdepositionOOIegg
The Acid Deposition Research Program is funded and adnninistered by the Province of
Alberta, the Canadian Petroleunn Association, Alberta's electrical utilities and the t^ANAPlANA
Energy Resources Conservation Board. ^
A distinctive feature of the ADRP is the development and funding of research in twyj^H22
major areas, biophysical and human health.
Acid Deposition Research Program - Members Committee
R.L. Findlay (Co-Chairman)
Manager, Environmental Affairs
AMOCO Canada Petroleum Co. Ltd.
Ken Smith (Co-Chairman)
Assistant Deputy Minister
Alberta Environment
Dr. John Railton
Manager, Environmental Planning
TransAlta Utilities
H. P. Sims
Director, Research Management Division
Alberta Environment
Ed Brushett
Manager, Environmental Protection
Energy Resources Conservation Board
Cornells (Casey) G. Van Teeling
Senior Manager, Research Management Division
Alberta Environment
E. A. Collom
Director of Environmental Health
Alberta Community and Occupational
Health
Program Manager: Dr. Ron Wallace
Communications Co-ordinator: Jean L, Andryiszyn
Doug Bruchett
Manager, Environment and Socio-Economic
Development
Canadian Petroleum Association
Scientific Advisory Board
Biophysical Research
Dr. Sagar V. Krupa (Chairman)
University of Minnesota
Department of Plant Pathology
Dr. C. M. Bhumralkar
National Oceanographic
and Atmospheric Administration (NOAA)
Dr. Herbert C. Jones
Tennessee Valley Authority
Fisheries and Aquatic Ecology Branch
Dr. Ron Kickert
Consultant (in modelling), Oregon
Dr. H. M. Liljestrand
University of Texas
Department of Civil Engineering
Dr. James P. Lodge
Consultant in Atmospheric Chemistry
and Editor, Atmospheric Environment
Dr. Douglas P. Ormrod
University of Guelph
Department of Horticultural Science
Dr. Carl L. Schofield
Cornell University
Department of Natural Resources
Dr. Robert K. Stevens
U.S. Environmental Protection Agency
Dr. M. Ali Tabatabai
Iowa State University
Department of Agronomy
Dr. T. Craig Weidensaul
Ohio State University
Ohio Agricultural Research
and Development Center
Biophysical Research Prime Contractor:
Kananaskis Centre for Environmental Research
The University of Calgary
Principal Investigator: Dr. Allan H. Legge
Acid Deposition Research Program, 3800, 150 Sixth Avenue S.W., Calgary, Alberta, Canada, T2P 3Y7
Acid
Deposition
Research Program
Reports in this set are available at no charge from: Program Manager
Acid Deposition Research Program
3800, 150 Sixth Avenue S.W.
Calgary, Alberta T2P 3Y7
World Literature Reviews
Report Number
□ ADRP-B-01-87
□ ADRP-B-02-87
□ ADRP-B-03-87
□ ADRP-B-04-87
□ ADRP-B-05-87
□ ADRP-B-06-87
□ ADRP-B-07-87
□ ADRP-B-08-87
□ ADRP-B-09-87
□ ADRP-B-10-87
References
Telang, S. A. 1987.
Surface Water Acidification Literature Review. Prep for the Acid Deposition Research Program by the Kananaskis Centre for Environmental Research, The University
of Calgary, Calgary, Alberta, Canada, 123 pp.
ISBN 0-921625-03-0 (Volume 1)
ISBN 0-921625-02-2 (Set of 11)
Visser, S., and Danielson, R. M., and Parr, J. F. 1987.
Effects of Acid-Forming Emissions on Soil Microorganisms and Microbially-Mediated Processes. Prep for the Acid Deposition Research Program by the Kananaskis
Centre for Environmental Research, The University of Calgary, Calgary, Alberta, Canada and U.S. Department of Agriculture, Beltsville, Maryland, U.S.A. 86 pp.
ISBN 0-921625-04-9 (Volume 2)
ISBN 0-921625-02-2 (Set of 11)
Krouse, H. R. 1987.
Environmental Sulphur Isotope Studies in Alberta: A Review. Prep for the Acid Deposition Research Program by the Department of Physics, The University
of Calgary, Calgary, Alberta. Canada. 89 pp.
ISBN-0-921625-05-7 (Volume 3)
ISBN-0-921625-0^ ^ w>-t of 11)
Laishley, E. J. and Bryant R. 1987.
Critical Review of Inorganic Sulphur Microbiology with Particular Reference to Alberta Soils. Prep for the Acid Deposition Research Program by the Department
of Biology, The University of Calgary, Calgary, Alberta, Canada. 56 pp.
ISBN 0-921625-06-5 (Volume 4)
ISBN 0-921625-02-2 (Set of 11)
Turchenek, L. W. and Abboud, S. A. and Thomas, C. J. and,Fessenden, R. J. and Holowaychuk, N. 1987.
Effects of Acid Deposition on Soils in Alberta. Prep for the Acid Deposition Research Program by the Alberta Research Council, Edmonton, Alberta, Canada. 202 pp.
ISBN 0-921625-07-3 (Volume 5)
ISBN 0-921625-02-2 (Set of 11)
Jaques, D. R. 1987.
Major Biophysical Components of Alberta. Prep for the Acid Deposition Research Program by Ecosat Geobotanical Surveys Inc 108 pp.
ISBN 0-921625-08-1 (Volume 6)
ISBN 0-921625-02-2 (Set of 11)
Campbell, K. W. 1987.
Pollutant Exposure and Response Relationships: A Literature Review. Geological and Hydrogeological Aspects. Prep for the Acid Deposition Research Program
by Subsurface Technologies and Instrumentation Limited, Calgary, Alberta, Canada. 152 pp.-l-maps.
ISBN 0-921625-09-X (Volume 7)
ISBN 0-921625-02-2 (Set of 11)
Torn, M. S. and Degrange, J. E. and Shinn, J. H. 1987.
The Effects of Acidic Deposition on Alberta Agriculture: A Review. Prep for the Acid Deposition Research Program by the Environmental Sciences Division,
Lawrence Livermore National Laboratory. 160 pp.
ISBN 0-921625-10-3 (Volume 8)
ISBN 0-921625-02-2 (Set of 11)
Mayo, J. M. 1987.
The Effects of Acid Deposition on Forests. Prep for the Acid Deposition Research Program by the Department of Biology, Emporia State University. 74 pp.
ISBN 0-921625-11-1 (Volume 9)
ISBN 0-921625-02-2 (Set of 11)
Krupa, S. V. and Kickert, R. N. 1987.
An Analysis of Numerical Models of Air Pollutant Exposure and Vegetation Response. Prep for the Acid Deposition Research Program by the Department
of Plant Pathology, University of Minnesota, St. Paul, Minnesota, U.S.A., and Consultant, Corvallis, Oregon, U.S.A. 113 pp.
ISBN 0-921625-12-X (Volume 10)
ISBN 0-921625-02-2 (Set of 11)
□ ADRP-B-11-87 Legge, A. H. and Crowther, R. A. 1987.
Acidic Deposition and the Environment: A Literature Overview. Prep for the Acid Deposition Research Program by the Kananaskis Centre for Environmental
Research, the University of Calgary, Calgary, Alberta, Canada. 235 pp.
ISBN 0-921625-13-8 (Volume 11)
ISBN 0-921625-02-2 (Set of 11)
ACIDIC DEPOSITION AND THE ENVIRONMENT:
A LITERATURE OVERVIEW
by
Allan H. Legge,
Kananaskis Centre for Environmental Research
The University of Calgary
Calgary, Alberta
and
R.A. Crowther
Aquatic Resource Management Limited
Calgary, Alberta
for subm'ission to
Acid Deposition Research Program
3860, 150 - 6 Avenue S.W.
Calgary, Alberta T2P 3Y7
November 1987
This publication may be cited as:
Legge, A,H. and Crowther, R,A. 1987.
Acidic Deposition and the Environment: A Literature Overview.
Prep for the Acid Deposition Research Program by Kananaskis
Centre for Environmental Research, The University of Calgary,
Calgary, Alberta, and Aquatic Resource Management Limited,
Calgary, Alberta. 235 pp.
ISBN 0-921625-13-8 (Volume 11)
ISBN 0-921625-02-2 (Set of 11)
i
PREFACE
The Alberta Government/Industry Acid Deposition Research Program (ADRP) is a
scientific investigation designed to answer specific questions regarding the environ-
mental effects of acidic and acid forming substances on the ecosystems of Alberta in
both the short and longer terms. As part of this study, a detailed and comprehensive
literature review of each of the identified critical areas of ecosystem concern was
completed. The following document is intended to provide an overview on the key
findings of each of these documents in both a world (Part I) and an Alberta (Part II)
context.
The documents that form the basis of this synthesis are:
ADRP-B-01/87: SURFACE WATER ACIDIFICATION LITERATURE REVIEW. Prep, for the
Acid Deposition Research Program by S.A. Telang, Kananaskis
Centre for Environmental Research, The University of Calgary,
Alberta, Canada. 132 pp. ISBN 0-921625-03-0 (Volume 1).
ADRP-B-02/87: EFFECTS OF ACID-FORMING EMISSIONS ON SOIL MICROORGANISMS AND
MICROBIALLY-MEDIATED PROCESSES. Prep, for the Acid Deposition
Research Program by S. Visser, R.M. Danielson, and J.F. Parr,
Kananaskis Centre for Environmental Research, The University of
Calgary, and U.S. Department of Agriculture, Beltsville,
Maryland. 86 pp. ISBN 0-921625-04-9 (Volume 2).
ADRP-B-03/87: ENVIRONMENTAL SULPHUR ISOTOPE STUDIES IN ALBERTA: A REVIEW.
Prep, for the Acid Deposition Research Program by H.R. Krouse,
Department of Physics, The University of Calgary, Calgary,
Alberta, Canada. 89 pp. ISBN 0-92165-06-5 (Volume 3).
ADRP-B-04/87: CRITICAL REVIEW OF INORGANIC SULPHUR MICROBIOLOGY WITH PARTICU-
LAR REFERENCE TO ALBERTA SOILS. Prep, for the Acid Deposition
Research Program by E.J. Laishley and R. Bryant, Department of
Biology, The University of Calgary, Calgary, Alberta, Canada.
50 pp. ISBN 0-921625-06-5 (Volume 4).
ADRP-B-05/87: EFFECTS OF ACID DEPOSITION ON SOILS IN ALBERTA. Prep, for the
Acid Deposition Research Program by L.W. Turchenek, S.A. Abboud,
C.J. Tomas, R.J. Fessenden, and N. Holowaychuk, Alberta Research
Council, Edmonton, Alberta. 202 pp. ISBN 0-921625-07-3
(Volume 5).
ADRP-B-06/87: MAJOR BIOPHYSICAL COMPONENTS OF ALBERTA. Prep, for the Acid
Deposition Research Program by D.R. Jaques, Ecosat Geobotanical
Surveys Inc., North Vancouver, British Columbia, Canada. 101
pp. + 4 maps. ISBN 0-921625-08-1 (Volume 6).
ADRP-B-07/87: POLLUTANT EXPOSURE AND RESPONSE RELATIONSHIPS: A LITERATURE
REVIEW. GEOLOGICAL AND HYDROGEOLOGICAL ASPECTS. Prep, for the
Acid Deposition Research Program by K.W. Campbell, Subsurface
Technologies and Instrumentation Limited, Calgary, Alberta,
Canada. 151 pp. + 2 maps. ISBN 0-921 625-09-X (Volume 7).
ADRP-B-08/87: THE EFFECTS OF ACIDIC DEPOSITION ON ALBERTA AGRICULTURE. Prep.
for the Acid Deposition Research Program by M.S. Torn, J.E.
Degrange, and J.H. Shinn, Lawrence Livermore National Labora-
tory, California, USA. 160 pp. ISBN 0-921625-10-1 (Volume 8).
ii
ADRP-B-09/87: THE EFFECTS OF ACID DEPOSITION ON FORESTS. Prep, for the Acid
Deposition Research Program by J.M. Mayo, Department of Biology,
Emporia State University, Emporia, Kansas, USA and the Kanan-
askis Centre for Environmental Research, The University of
Calgary, Alberta, Canada. 74 pp. ISBN 0-921625-11-1 (Volume 9).
ADRP-B-lO/87: AN ANALYSIS OF NUMERICAL MODELS OF AIR POLLUTANT EXPOSURE AND
VEGETATION RESPONSE. Prep, for the Acid Deposition Research
Program by S.V. Krupa and R.N. Kickert, Department of Plant
Pathology, University of Minnesota, St. Paul, Minnesota, USA
and Consultant, Corvallis, Oregon, USA. 113 pp. ISBN 0-921625-
12-X (Volume 10).
iii
TABLE OF CONTENTS
Page
PREFACE i
TABLE OF CONTENTS iii
LIST OF TABLES vi
LIST OF FIGURES viii
ACKNOWLEDGEMENTS ix
1. INTRODUCTION TO ATMOSPHERIC CHEMISTRY AND ACIDIC DEPOSITION PROCESSES . . 1
1.1 Atmospheric Processes 1
1.2 Deposition Processes 2
1.3 Chemistry of Precipitation 3
1.4 Wet Deposition in Alberta 7
1.5 Introduction to Atmospheric Chemistry and Acidic Deposition Processes:
Literature Cited 12
2. EFFECTS OF ACIDIC DEPOSITION ON FORESTS 17
2.1 Introduction 17
2.2 Forest Concerns Related to Acidic Deposition 18
2.2.1 Forest Decline Phenomenon 18
2.3 Direct Effects of Acidic Deposition on Forests 19
2.4 Indirect Effects of Acidic Deposition on Forests 19
2.4.1 Canopy-Pollutant Interactions 19
2.5 Interactive Effects of Acidic Deposition on Forests 32
2.5.1 Interactive Effects on Forest Nutrition and Growth 32
2.5.2 Timber Harvesting and Acidic Deposition 33
2.5.3 Effects of Acidic Deposition on Tree Reproduction 34
2.6 Effects of Acidic Deposition on Plant Communities 34
2.7 Effects of Acidic "Deposition on Forests: Literature Cited 35
3. ACIDIC DEPOSITION EFFECTS ON AGRICULTURE 43
3.1 Effects of Acidic Precipitation on Crops 43
3.2 Foliar Injury 43
3.3 Sensitivity of Plants to Foliar Injury Caused by Wet Acidic Deposition . . 47
3.3.1 Direct Foliar Effects of Wet Acidic Deposition 51
3.3.1.1 Foliar Fertilization 51
3.3.1.2 Foliar Buffering 52
3.3.1.3 Foliar Leaching 52
3.3.1.4 Foliar Nutrient Content " 53
3.4 Effects of Wet Acidic Deposition on Plant Growth 54
3.5 Effects of Wet Acidic Deposition on Plant Reproduction 57
3.6 Effects of Dry Deposition on Agricultural Crops 57
3.6.1 Physiological Effects of Dry Deposition 58
3.6.1.1 Sulphur Dioxide Effects on Stomata 58
3.6.1.2 Sulphur Dioxide Effects on Photosynthesis 59
3.6.1.3 Sulphur Dioxide Effects on Respiration 59
3.6.1.4 Nitrogen Oxide Effects on Stomata and Transpiration 59
3.6.1.5 Nitrogen Oxide Effects on Photosynthesis 60
3.6.1.6 Nitrogen Oxide Effects on Respiration 60
3.6.1.7 Ozone Effects on Stomata, Transpiration, and Photosynthesis 60
3.6.2 Foliar Effects of Dry Deposition 60
3.6.2.1 Foliar Effects of Sulphur Dioxide 61
3.6.2.2 Foliar Effects of Nitrogen Oxide 61
3.6.2.3 Foliar Effects of Ozone 66
3.6.3 Growth and Yield Effects of Dry Deposition 66
3.6.3.1 Effects of Sulphur Dioxide on Growth and Yield 66
3.6.3.2 Effects of Nitrogen Oxide on Growth and Yield 67
3.6.3.3 Effects of Ozone on Growth and Yield 67
3.6.4 Effects of Dry Deposition on Plant Reproduction 73
3.7 Effects of Mixtures of Gaseous Pollutants on Crops 73
3.7.1 Combined Effects of Sulphur Dioxide and Ozone 78
3.7.2 Combined Effects of Sulphur Dioxide and Nitrogen Dioxide 78
3.7.3 Combined Effects of Nitrogen Dioxide and Ozone 79
3.7.4 Combined Effects of Sulphur Dioxide, Nitrogen Dioxide, and Ozone 80
3.8 Combined Effects of Dry and Wet Deposition 80
i V
TABLE OF CONTENTS (continued)
Page
3.9 Effects of Acidic Deposition on Plant-Soil Interactions 80
3.9.1 Effects of an Acidified Soil Environment on Plants 81
3.9.2 Effects of Altered Soil Acidity on Soil Organism-Plant Interactions ... 81
3.9.2.1 Effects of Acidic Deposition on Plant-Microbe Interactions 85
3.9.2.2 Effects on the Plant as Host Organism 85
3.9.2.3 Effects on Viruses, Fungi, and Bacteria 85
3.9.2.4 Effects on Insect-Plant Relationships 85
3.10 Acidic Deposition Effects on Agriculture: Literature Cited 92
4. NUMERICAL MODELS OF AIR POLLUTANT EXPOSURE AND VEGETATION RESPONSE .... 105
4.1 Types of Models 105
4.2 Acute versus Chronic Exposure .... 105
4.3 Characteristics of Ambient Air Quality 106
4.4 The Concept of Pollutant Dose 107
4.5 Mathematical Models for Characterizing Plant Response to Air
Pollutant Stress 108
4.5.1 Acute Pollutant Exposure and Plant Response Models 108
4.5.2 Chronic Pollutant Exposure and Plant Response Models 108
4.6 Numerical Models of Pollutant Exposure and Vegetation Response:
Literature Cited Ill
5. EFFECTS OF ACIDIC DEPOSITION ON SOILS 115
5.1 Acid-Base System in Soils 115
5.2 Soil Reactions 115
5.3 Total Soil Acidity 118
5.4 Cation Exchange and Soil Acidity 118
5.5 Base Saturation 119
5.6 Natural Acidification of Soils 119
5.6.1 Acidification in Soil Genesis 119
5.6.2 Natural Sources of Soil Acidity 120
5.6.2.1 Organic Matter 120
5.6.2.2 Leaching and Weathering 123
5.7 Influences of Soil Acidity and Acidification on Soil Properties 125
5.7.1 Organic Matter 125
5.7.2 Soil Cations and Leaching 128
5.7.3 Soil Anions 130
5.7.4 Availability of Nutrients and Toxic Metals 130
5.8 Effects of Anthropogenic Sources of Acidity 131
5.8.1 Nitrogenous Fertilizers 131
5.8.2 Atmospheric Deposition 132
5.9 Summary . 133
5.10 Effects of Acidic Deposition on Soils: Literature Cited 137
6. EFFECTS OF ACIDIC DEPOSITION ON SOIL MICROORGANISMS AND MICROBIALLY
MEDIATED PROCESSES ... 143
6.1 Introduction 143
6.2 General Effects of Acidic Deposition on Soil Microbes 143
6.2.1 Influence of Soil Acidity on Microbial Communities 144
6.2.2 Summary of Acidic Deposition Effects on Microbial Processes 145
6.3 Acidic Deposition and Inorganic Sulphur Microbiology 147
6.3.1 Oxidation Reactions 148
6.3.2 Heterotrophic Microorganisms 151
6.3.3 Reduction Reactions 151
6.4 Ecological and Economic Effects of Microbial Inorganic Sulphur Oxidation
and Reduction 153
6.4.1 Oxidation of Metal Sulphides in Soil 153
6.4.2 Phototrophic Sulphur Bacteria 155
6.5 Factors Affecting the Microbial Oxidation of Sulphur 155
6.5.1 Sulphur 155
6.5.2 Soil Environment and Its Effects on Sulphur Microbiology 156
6.6 Effects of Acidic Deposition on Soil Microorganisms and Microbially
Mediated Processes: Literature Cited 158
V
TABLE OF CONTENTS (concluded)
Page
7. EFFECTS OF ACIDIC DEPOSITION ON GEOLOGY AND HYDROGEOLOGY 161
7.1 Groundwater Hydrology 161
7.2 Hydrogeological Neutralization Processes 161
7.3 Evidence of Groundwater Acidification 163
7.4 Effects of Acidic Deposition on Major Cations and Anions in Groundwater . 163
7.5 Effects of Acidic Deposition on Metals in Groundwater 164
7.6 Prediction of Acidic Deposition Effects on Groundwater 164
7.6.1 Sensitivity Analysis 164
7.6.2 Modelling 164
7.6.3 Human Impacts 165
7.7 Effects of Acid Deposition on Geology and Hydrogeology: Literature Cited . 166
8. EFFECTS OF ACIDIC DEPOSITION ON SURFACE WATER ACIDIFICATION 169
8.1 Determination of Acidity in Surface Waters 169
8.2 Sensitive Waters 170
8.3 Watershed Characteristics Determining Surface Water Susceptibility to
Acidification 170
8.3.1 Major Determining Factors of Surface Water Acidity 171
8.3.1.1 Forest Canopy 171
8.3.1 .2 Bedrock Geology 171
8.3.1.3 Soil Type and Depth 171
8.3.1.4 Topography and Watershed-to-Lake Ratio 175
8.3.1.5 Watershed Vegetation and Land Use 175
8.3.1.6 Surface Water Quality 175
8.3.1.7 Climate and Meteorological Conditions . 176
8.4 Acidic Waters and Their Reaction Products 176
8.5 Precipitation Quantity and Quality as Factors in Surface Water
Acidification 178
8.6 Potential Sources of Acidification of Surface Waters 181
8.7 Trends in Surface Water Acidification in North America 181
8.8 Effects of Acidic Deposition on Aquatic Biota 184
8.9 Models of Freshwater Acidification 185
8.9.1 ILWAS Model 185
8.10 Effects of Acidic Deposition on Surface Water Acidification:
Literature Cited 195
9. ACIDIC DEPOSITION IN THE ALBERTA CONTEXT 203
9.1 Major Biophysical Components of Alberta 203
9.2 Acidic Deposition and Alberta Forests 203
9.3 Acidic Deposition and Alberta Agriculture 206
9.3.1 Wet Deposition Effects on Agriculture 206
9.3.2 Dry Deposition Effects on Agriculture 207
9.4 Alberta Soils Sensitive to Acidic Deposition 211
9.4.1 Soil Sensitivity and Mapping 211
9.4.2 Chernozemic Soil Impacts 213
9.4.3 Solonetzic Soil Impacts 213
9.4.4 Luvisolic Soil Impacts 215
9.4.5 Brunisolic and Podzolic Soil Impacts 217
9.4.6 Organic and Organic Cryosolic Soils 217
9.4.7 Gleysolic Soil Impacts 218
9.4.8 Regosolic Soil Impacts 218
9.4.9 Impacts on Rocklands and Rough-Broken Lands 218
9.5 Effects of Acidic Deposition on Soil Microorganisms and Processes .... 218
9.6 Sulphur Microbiology in the Alberta Context 219
9.7 Surface Water Acidification Studies in Alberta 222
9.8 Alberta Hydrogeology and Geology 225
9.8.1 Geological Information Bases for Alberta 225
9.8.1.1 Bedrock Geology 225
9.8.1.2 Surficial Geology 225
9.8.2 Hydrogeology Resource Inventory 226
9.9 Sulphur Isotope Studies in Alberta 226
9.10 Acidic Deposition in the Alberta Context: Literature Cited 231
vi
LIST OF TABLES
Page
1. Some inorganic ions important in precipitation chemistry 4
2. Wet deposition in Alberta (1978-1984) 8
3. Wet deposition of H+, SO42-, and NOa" (kg ha'^ y-^) in Alberta
and at selected Canadian stations from 1978 to 1982 9
4. Modelled dry and dry-wet sulphate deposition ratios for Alberta
sites (1982) 10
5. References to various effects of acidic deposition on soils, plants
forests, and ecosystems 20
6. The effects of pollutants on stomatal diffusive resistance and
water status 21
7. The effects of pollutants on photosynthesis and carbon allocation .... 23
8. Biochemical effects of pollutants 26
9. Effect of pollutants on reproductive biology 28
10. Potential effects of acidic precipitation on vegetation 45
11. Visible foliar injury resulting from simulated wet acidic deposition:
pH threshold 48
12. Effect of simulated acidic rain on marketable yield of roots and shoots . 55
13. Threshold sulphur dioxide concentrations (ppm) causing foliar injury to
various agricultural species 62
14. Agricultural species sensitive to sulphur dioxide 64
15. Suggested susceptibility of various agricultural species which occur
in Alberta to a combination of nitrogen dioxide and nitric oxide 65
16. Yields of two field crops grown in different concentrations of
sulphur dioxide 68
17. Yield of various crops in field plots exposed to sulphur dioxide 69
18. Effects of sulphur dioxide on cultivars of hard red spring wheat
(HRS) and soft white winter wheat (SWW) 70
19. Effects of acute ozone exposure on growth and yield of agricultural
crops 71
20. Effects of long-term controlled ozone exposures on growth, yield, and
foliar injury of various agricultural species 74
21. Toxic concentration of copper, nickel, or zinc in leaf tissue 83
22. Plant sensitivity to acid-induced changes in the soil environment .... 83
23. Recommended crops for soils with varying acidity in Great Britain .... 84
24. Effect of drop of 0.1 unit in soil pH on barley and alfalfa yield .... 84
25. Effect of pollutants on plant -pathogen interactions 86
26. Simulated acid rain-fungal life cycle interaction 90
27. Summary of the assessment of applicability of the statistical
(empirical) models reviewed by Krupa and Kickert (1987) 109
vii
LIST OF TABLES (concluded)
Page
28. Summary of the assessment of applicability of the mechanistic
(process oriented) models reviewed by Krupa and Kickert (1987) 110
29. Summary of proton producing and consuming processes 124
30. Summary of the potential impact of acidic deposition on soils 134
31. Estimated amounts of cultivated land in different ranges of soil pH on
the Great Plains 136
32. The effect of pH on soil microorganisms 146
33. Chemical reactions of the thiobacilli 150
34. Watershed characteristics that influence surface water susceptibility to
acidification 172
35. Surface water acidification studies reviewed by Telang (1987) 182
36. Lower pH limits for various groups of organisms in naturally acidic waters 186
37. Effects of increasing acidity on aquatic ecosystems 188
38. Alberta research references on the effects of pollutants on forest
ecosystems 204
39. Effects of ambient sulphur dioxide on yield of various agricultural
species 208
40. Susceptibility of various agricultural species which occur in Alberta
to nitrogen dioxide 209
41. Agricultural crops grown in Alberta which are known to be relatively
sensitive to ozone 210
42. Areas of the soil orders in Alberta 212
43. Chernozemic soils sensitive to acidic deposition 214
44. Modelled predictions (Bloom and Grigal 1985) for soil pH responses to
acid inputs 216
45. Indicator parameters used to classify the sensitivity to acidification
of Alberta lakes 223
vi i i
LIST OF FIGURES
Page
1. Relationship between soil pH and activity of microorganisms and
availability of plant nutrients 82
2. Major sources and sinks of acidity in soil 121
3. Reduction in cation exchange capacity of organic matter and clay with
decrease in soil pH 126
4. Global sulphur cycle 149
5. Oxidation of lactic acid and the di ssimi latory reduction of sulphate
by Desulf ovibrio sp 152
6. Direct and indirect oxidation mechanisms for pyrite oxidation 154
7. Schematic diagram of the hydrologic cycle 162
8. Precipitation pathways to a lake 173
9. Lateral flow of water from different soil layers in determining lake
water pH 174
10, Chemical species associated with water flow paths to a lake 174
11. Location of soil testing areas in Alberta, and the percentage of
cultivated soil with a pH of 6.0 or less for each area 221
ix
ACKNOWLEDGEMENTS
The authors gratefully acknowledge the financial assistance of the Alberta
Government/Industry Acid Deposition Research Program (ADRP) for the preparation of this
literature overview. The authors extend their appreciation to the Co-chairmen and the
Members' Committee of the ADRP and the ADRP Program Manager, Dr. R.R. Wallace. The
critical review of this document by Dr. S.V. Krupa and other members of the Scientific
Advisory Board of the ADRP was most helpful and very much appreciated. The assistance of
Ms. Jean Andryiszyn of Francis, Williams & Johnson Limited in the publication of this
document, as well as the other documents in this series, is appreciated. The authors
wish to thank Delia Patton and Lynn Ewing for their assistance in typing this overview.
Finally, the assistance, dedication, and skill of Linda Jones in the final preparation of
this report is gratefully acknowledged.
X
PART I.
GENERAL OVERVIEW AND WORLD PERSPECTIVE:
ACIDIC DEPOSITION AND ECOSYSTEM EFFECTS
1
1 . INTRODUCTION TO ATMOSPHERIC CHEMISTRY AND ACIDIC DEPOSITION PROCESSES
The short- and long-term observed and/or predicted, adverse impacts of air
pollutants on terrestrial and aquatic ecosystems are of utmost concern at this time
(U.S. National Research Council 1983; U.S. Environmental Protection Agency 1983). Air
pollutants occur as gases, vapours, and particulate matter (both dry and wet). Once
pollutants are emitted, the atmosphere serves as a medium for their dilution, transport,
chemical reaction, and deposition. These processes are governed by the physical and
chemical properties of the pollutant emitted, its reactivity with the other constituents
in the atmosphere, and by the meteorological conditions. In the end, primary and
secondary pollutants are transferred from the atmosphere to surfaces (crops, forests,
soils, surface waters, and materials) by dry and wet deposition.
Dry deposition may be defined as the direct collection of gaseous and par-
ticulate species on land and surface waters (Garland 1978). On the other hand, wet
deposition comprises the incorporation of the pollutant in cloud droplets (rainout) and
removal by falling precipitation (washout). The relative importance of the two deposition
processes is known to vary in time and space (U.S. National Research Council 1983).
Thus, any observed or predicted receptor responses are considered to be due to the joint
action of both dry and wet deposition (Legge and Krupa 1986).
What follows in the subsequent sections of this "Introduction" was essentially
extracted from Krupa et al. (1987a).
1.1 ATMOSPHERIC PROCESSES
The occurrence of "acidic precipitation" is of much concern. The acidity of
precipitation, particularly in the industrialized areas, is considered to be due to the
presence of strong mineral acids in the atmosphere (Hutchinson and Havas 1980; Inter-
national Electric Research Exchange 1981; U.S. National Research Council 1983; U.S.
Environmental Protection Agency 1983; and U.S. National Research Council 1986).
The combustion of sulphur-containing fossil fuels leads to the emission of
sulphur dioxide (SO2). Once emitted, depending upon meteorological and other condi-
tions, a highly variable portion of the SO2 is continuously (except during precipitation
events) deposited on to surfaces by dry deposition. Particularly during the daylight
hours, SO2 is also oxidized in the emission plume and ambient atmosphere to sulphuric
acid (H2SO4), aerosols, or sulphates by reactions occurring in the gas phase, in the
liquid phase, on the surfaces of solids, or through combinations of all three (Finlayson-
Pitts and Pitts 1986). For details of the SO2 oxidation mechanisms, the reader is
referred to Finlayson-Pitts and Pitts (1986), Hidy and Mueller (1986), and U.S. National
Research Council (1983).
In power plant plumes, SO2 oxidation rates of up to 4% h~^ have been reported
(Husar et al. 1978). Often such rates are much higher if the plume passes through clouds
or fog banks (Eatough et al. 1984). Similarly, the rates of production of sulphate
(S04^ ) from SO2 are much higher during the summer compared with the winter (Richards et
al. 1981). According to Gillani (1978) and Forrest et al. (1981), noontime SO2 conversion
rates in a power plant plume were 1-4% h~^ compared with night-time rates of <0.5% h~^.
However, significant S04^" production (4.5 to 10.8% h~^) can occur at night-time if
clouds are a contributing factor in the SO2 conversion (Cass and Shair 1984). For
2
details of the suggested mechanisms for the aqueous oxidation of SO2, the reader is
referred to Graedel and Goldberg (1983), Bielski et al. (1985), and Graedel et al.
(1986).
Compared with SO2, the oxidation of the oxides of nitrogen (NOx) in power
plant plumes and in ambient air is relatively less understood. In power plant plumes,
rates of conversion of NOx from roughly 0.2 to 12% h ^ have been observed, with the
rates being much greater during midday than at night (Hegg and Hobbs 1979; Richards
et al. 1981). The products of NOx oxidation appear to be peroxyacetyl nitrate (PAN) and
nitric acid (HNOa), with a lesser amount of particulate nitrates (NO3 ). In urban plumes,
rates of conversion of NOx of <5% h ^ to 24% h ^ have been reported (Chang et al. 1979;
Spicer 1982a, b).
In the atmosphere, H2SO4 and HNO3 differ in their physical and chemical
behaviour. Nitric acid is more volatile and thus, significant concentrations of that
substance can exist in the gas phase. On the other hand, H2SO4 has a low vapour
pressure under ambient conditions and exists in the fine particle phase (<2.0 ym)
(Whitby 1978; Roedel 1979). These particles, in addition to causing visibility degrada-
tion, can also act as cloud condensation nuclei (Husar et al. 1978).
Both H2SO4 and HNOa can react with bases present in the atmosphere to form
salts. For example, H2SO4 can react rapidly with ammonia (NHa) in the atmosphere to form
ammonium acid sulphate (NH4HSO4), letovicite ( (NH4) 3H(S04) 2) , and ammonium sulphate
((NH4)2S04). Similarly, HNOa can react with NHs to form ammonium nitrate (NH4NO3).
Because of the characteristics of the equilibrium between HNOa, NHa, and NH4NO3, HNO3 can
revolatilize relatively easily even after forming the ammonium salt ( Finlayson-Pitts and
Pitts 1986). No such analogous physical and chemical changes exist for H2SO4. In
addition to the ammonium salts, H2SO4 and HNOs can readily form salts with other cations,
such as Ca^^, Mg^^, and so forth.
1.2 DEPOSITION PROCESSES
According to Garland (1978) deposition processes limit the lifetime of sulphur
and other pollutants in the atmosphere, control the distance travelled before deposition,
and limit their atmospheric concentrations. Therefore, an understanding of such proces-
ses is essential for a proper assessment of the environmental significance of natural
and man-made emissions of sulphur and other pollutants.
Dry deposition leads to the direct collection of gases, vapours, and particles
on land and water surfaces. This pollutant transfer process includes diffusion, Brownian
motion, interception, impaction, and sedimentation (U.S. National Research Council 1983;
Legge and Krupa 1986; and Voldner et al. 1986). The rate at which these processes
transfer pollutants from the air to exposed surfaces is controlled by a wide range of
chemical, physical, and biological factors which vary in their relative importance
according to the nature and state of the surface, the characteristics of the pollutant,
and the state of the atmosphere (U.S. National Research Council 1983). The complexity
of the individual processes involved and the variety of possible interactions among them
combine to prohibit easy generalization; nevertheless, a "deposition velocity" v^ ,
analogous to a gravitational falling speed, is of considerable use. In practice,
knowledge of v^ enables fluxes, F, to be estimated from air concentrations, C, as the
simple product, v. • C (U.S. National Research Council 1983).
3
For more information on the dry deposition of air pollutants, the reader is
referred to Garland (1978), International Electric Research Exchange (1981), U.S.
National Research Council (1983), and Chamberlain (1986).
Both SO2 and S04^ can contribute significantly to the dissolved sulphur in
rain. The contribution of S04^ appears inevitable, since SO*^ particles serve as cloud
condensation nuclei. On the other hand, the incorporation of SO2 may be suppressed if
the condensation nuclei are acidic. Dana et al. (1975) imply no more than 3% deposition
in rain from a power plant plume in the first 10 km. Larson et al. (1975) deduced that
only 8% of the sulphur emitted from a smelter while rain was falling was deposited within
60 km. Garland (1978) has summarized the information on mechanisms contributing to
sulphur in rainwater.
In addition to the rainout of condensation nuclei, several other physical
processes may contribute to sulphate in rain. Dif fusiophoresis and Brownian diffusion
may result in the collection of small particles on to the cloud droplets and raindrops
may further collect particles by impaction, interception, or diffusion. According to
Garland (1978) only the rainout of condensation nuclei appears capable of explaining the
concentrations of several mg L ^ of SOa^~ observed in practice.
The washout of large particles by raindrops may make a significant contribution,
but this fraction of the aerosol will be exhausted by the first few millimetres of rain
and may, therefore, account for the enhancement in sulphate concentration observed at
the beginning of some periods of rain (Meurrens 1974; Pratt et al . 1983).
Diffusion and interception may be of greater significance in snow because of
the larger surface area of the precipitation elements. In addition, the concentration
of condensation nuclei collected in precipitation may be much reduced if distillation
from liquid to solid phase dominates the aggregation of cloud droplets in the growth of
snowflakes.
In summary, the probable contribution to acidity of dissolved SO2 is smaller
than the contribution due to the rainout of S04^~. However, oxidation of SO2 in
clouds can make a substantial contribution to the S04^~ in rain.
In contrast to S04^ , much less information has been published regarding
the removal mechanisms of nitrogen species by precipitation. There is some evidence for
the formation of HNOa in clouds and rainwater. Recently, both theory and experimental
evidence suggest that HNOa may be formed rapidly from a combined gas phase/liquid
phase process (U.S. National Research Council 1983). Although significant uncertainty
remains concerning the source of HNOa in clouds and rainwater, the limited evidence
currently available favours the probable importance of the formation of nitrogen pent-
oxide (N2O5) from nitrogen dioxide (NO2), followed by its reaction with water
droplets to form HNOa ( Finlayson-Pitts and Pitts 1986). It also appears that HNOa
can be effectively scavenged by precipitation.
1.3 THE CHEMISTRY OF PRECIPITATION
Because of the significant concern arising from the occurrence of "acidic
precipitation", numerous investigators have examined the qualitative and quantitative
aspects of precipitation chemistry in the last 15-20 years. For more details than those
presented in the following section, the reader is referred to Husar et al. (1978),
4
Chamberlain et al. (1981), U.S. National Research Council (1983), Teasley (1984), and
U.S. National Research Council (1986).
In comparison with seawater (Whitfield 1979), precipitation can be considered
as a highly unbuffered, dilute solution of organic and inorganic ions. Precipitation is
also composed of an insoluble fraction consisting of organic (e.g., pollen, pesticides,
and so forth) and inorganic (e.g., crustal) coarse (>2.0 ym size) particles (Krupa
et al. 1976). However, it is the soluble fraction which is of concern in the context of
"acidic precipitation". Table 1 lists some of the inorganic ions important in precipi-
tation chemistry. Ecological effects scientists have utilized the concentrations of
many of these ions together with precipitation depth to compute ion deposition (kg.ha"^)
in order to evaluate effects (U.S. Environmental Protection Agency 1983).
Table 1. Some inorganic ions important in precipitation chemistry.*
Cations
Anions
H+
NH4+
ci-
Na+
NO3-
K+
SO32-
Ca2+
SO42-
Mg2+
PO42-
COa^-
All ions are presented here in their completely dissociated states.
The reader should note, however, that various states of partial
dissociation are possible as well (e.g., HSOa", HCOa") .
Source: U.S. National Research Council (1983).
The pH of natural precipitation is often assumed to be regulated by the dissoci-
ation of dissolved carbon dioxide (CO2), thus having a value of 5.6. Precipitation pH
values below 5.6 are therefore assumed to be due to the addition of acidic components
(primarily related to S04^'' and NOa") by human activity (Garrels and Mackenzie 1971;
Likens and Bormann 1974; and Galloway et al. 1976).
According to some investigators, the acidity and concentrations of S04^
and NO3 and some other components in precipitation have increased in recent years
in certain geographic locations as a result of human activities (Cogbill and Likens
1974; Galloway et al. 1976; Martin and Barber 1977; and Likens and Butler 1981).
The data from Hubbard Brook, New Hampshire revealed several trends (Likens
et al. 1980) which were supported at least qualitatively by bulk deposition monitoring
data (but which had relatively unreliable quality control) from nine sites in New York
State (Miles and Yost 1982; Peters et al. 1982). These trends indicated that:
5
1. There has been a decrease in S04^ concentration since 1964 but an
increase in NOa concentration over the same time;
2. The annual pH of precipitation showed no long-term, significant change
from 1964 to 1977, although several short-term changes did occur;
3. A linear regression equation of data points from 1964 to 1977 indicated no
statistically significant trends in deposition; and
4. Recent changes in deposition correspond more with changes in NOa"
deposition than with S04^~ deposition, even though H2SO4 is the
dominant species at Hubbard Brook. The contribution of NOa" to total
acidity has been increasing, whereas that of S04^~ has been decreasing
(Galloway and Likens 1981). Year-to-year changes superimposed on the
long-term trend may be related to cl imatological influences (U.S. National
c Research Council 1983).
According to Hansen et al . (1981) available data are not of sufficient quantity
and quality to support any long-term trends in precipitation acidity change over the
past 50 years in the eastern United States. However, the observations do show that
precipitation is definitely acidic over this region, and is probably more acidic than
expected from natural baseline conditions.
Recently, Schertz and Hirsch (1985) performed a trend analysis (1978-1983) of
data from 19 sites of the NADP (National Atmospheric Deposition Program). They concluded
that 41% of the trends detected in the ion concentrations were downward trends, 4% were
upward trends, and 55% showed no trends at a = 0.2. The authors also concluded that
the two constituents of greatest interest in terms of human-generated emissions and
environmental effects, S04^" and NOa", showed only downward trends, and S04^~ showed the
largest decreases in concentration per year of all the ions tested.
On the other hand, Stensland et al. (1986) in their analysis of long-term
trends, derived the following conclusions:
1. The eastern half of the United States experiences concentrations of S04^~
and NOa" in precipitation that are, in general, greater by at least a fac-
tor of five than those in the remote areas of the world, indicating that
levels have increased by this amount in northeastern North America since
sometime before the 1950' s;
2. Data on the chemistry of precipitation before 1955 should not be used for
trend analysis;
3. Precipitation is currently more acidic in parts of the eastern United
States than it was in the mid-1950's or mid-1960's; however, the amount of
change and its mechanism are in dispute;
4. Precipitation S04^ concentrations and possibly acidity have increased
in the southeastern United States since the mid-1950's; and
5. In general, individual sites or groups of a few neighboring sites cannot
be assumed 'a priori' to provide regionally representative information;
regional representativeness must be demonstrated on a site-by-site basis.
6
Assuming that acidity has increased (at least in certain locations) and that
SO2 and NOx are responsible for most of the free acidity, a strong statistical rela-
tionship should be observed between and S04^ and/or N03 . Information to date
suggests that the degree of association is site specific and that proximity to sources of
SO2 and NOx, as well as to sources of other substances, may influence the chemical
nature of acidic substances in the atmosphere (Lefohn and Krupa 1984). At a given site,
differences in the meteorology between events may result in wide variations in the
measured acidity, concentrations of S04^~ and NOa", and correlations between these ions
(Pratt et al. 1984; Pratt and Krupa 1985). Sequeira (1982) found correlation coeffici-
ents in excess of 0.8 between S04^" and h"*" at Mauna Loa, Hawaii, somewhat lower
values at Monte Cimone, Italy, and 0.01 at Alamosa, Colorado. Barrie (1981) examined
summer data from eastern Canadian sites and found correlation coefficients (r^ val-
2- +
ues) between SO4 and H as high as 0.99 at some sites and as low as -0.39 at
other sites. Similarly, Pratt et al . (1983), examining four years of rainfall chemistry
at seven sites in a 600 km^ area in central Minnesota, found correlation coefficients
between S04^ and to vary from r^ values of 0.15 to 0.42, and between NOa and
h"^ from 0.06 to 0.62.
Kasina (1980) found no significant correlation between acidity and S04^~ in
southern Poland. On the other hand, Madsen (1981) found good correlations between
and excess S04^~ on the east coast of Florida during most of the months from late
1977 to late 1979. McNaughton (1981) found correlation coefficients in excess of 0.7
for all MAP3S/RAINE (1982) sites except Illinois, where the value was below 0.4.
The preceding discussion shows the complexity in generalizing the characteris-
tics of precipitation chemistry because of significant spatial variability. In addition,
it is well known that precipitation chemistry exhibits distinct temporal variability
including seasonality (Pratt and Krupa 1983; Dana and Easter 1987). For example,
Bowersox and de Pena (1980) concluded, by applying multiple linear regression analysis
for a central Pennsylvania site, that on average H2SO4 was the principal contributor
to concentration in rain, but that the acidity in snow was principally from HNOa.
According to Sequeira (1982), the pH of atmospheric precipitation at a given
location depends on the chemical nature and relative proportions of acids and bases in
the solution. Sequeira concluded that a pH of 5.6 may not be a reasonable reference
value for unpolluted precipitation. Charlson and Rodhe (1982) also questioned the
validity of using pH 5.6 as the background reference point, citing naturally occurring
acids as possibly responsible for low pH values of rain. They stated that consideration
of the natural cycling of water and sulphate through the atmosphere, precipitation rates,
and experimentally determined rates of S04^ scavenging indicates that average pH
values of approximately 5.0 would be expected in pristine locations in the absence of
basic materials. This value will vary considerably due to the variability in scavenging
efficiencies as well as geographic patchiness in the sulphur and hydrological cycles.
Thus, precipitation pH values might range from 4.5 to 5.6 due to these variabilities
alone (Charlson and Rodhe 1982). Recently, Lefohn and Krupa (1987) found that pH of
precipitation with minimum concentrations of S04^~ and NOa was in the range of 4.6
to 5.5 for the northeast United States.
7
An aspect of precipitation chemistry which has been largely ignored until
recently is the presence of organic acids (Krupa et al. 1976). Meyers and Hites (1982),
Kawamura and Kaplan (1983), Keene and Galloway (1984), Guiang et al. (1984), and Chapman
et al. (1986) have all shown the presence of organic acids in precipitation. Keene and
Galloway (1984) estimated that organic acids may contribute 16 to 35% of the volume
weighted free acidity in precipitation of North America. Krupa et al. (1987b) calculated
theoretical precipitation concentrations for Minnesota, based on the assumption
that all of the organic anions were present as the corresponding acids. A plot of the
calculated versus the measured h"*" concentrations showed poor correlation, yet the mean
calculated and the mean measured concentrations were nearly equal, on a yearly
basis. The weak organic acids could account for all of the deposited acidity.
Independent of these considerations of the complexity of precipitation chemis-
try, it is widely accepted that SOa^ and NOa form the basis for acidic precipitation
(U.S. National Research Council 1983). Statistical parameters such as the mean and
standard deviation, which can be interpreted unambiguously when the distribution of data
is normal or Gaussian, do not accurately characterize the observed distributions of
precipitation ions (Knapp et al. 1987). These distributions appear to be best described
by the mathematical functions of the Weibull family (Nosal and Krupa 1986).
Numerous investigators have used linear or linearizable statistical methods in
evaluating the relationships between the major inorganic ions in precipitation. The use
of such techniques assumes that: (1) the data are normally distributed and independent;
and (2) the residuals in the regression are normally distributed, independent, unbiased,
and homoscedastic (Snedecor and Cochran 1978). Recently, Nosal and Krupa (1986) showed
that the aforementioned assumptions are violated by the data on precipitation chemistry.
An additional implication of this finding is that precipitation data between sites cannot
be pooled or combined in performing parametric statistical analyses of either the
precipitation characteristics or acidic precipitation-receptor response relationships.
1.4 WET DEPOSITION IN ALBERTA
Since the early 1970's, several investigators have studied the precipitation
composition and wet deposition in Alberta (Nyborg et al. 1977; Caiazza et al. 1978; Klemm
and Gray 1982; Lau 1985; and Lau and Das 1985). Based on a study during 1973-1974,
Nyborg et al. (1977) concluded that rain and snow in Alberta were seldom acidic. Klemm
and Gray (1982), in their study during 1977-1978, found that less than 20% of the pH
values of precipitation in central Alberta were below 5.0 and none were below 4.0.
According to Lau and Das (1985) volume weighted average pH values of composite monthly
precipitation samples at 11 Alberta CANSAP (Canadian Network for Sampling Precipitation)
stations during 1978-1984 ranged between 5.17 and 6.06 (Table 2). Wet deposition of
H (kg ha~^ y~^) in Alberta during 1978-1982 was 24 times less compared with the values
for Ontario, Quebec, and Nova Scotia (Table 3). In a similar comparison, SOa^~
deposition in Alberta was 1.7 to 3.7 times less and NOa" deposition was 1.8 to 6.0
times less compared with the three eastern states (Table 3). Kociuba (1984) based on
modelling concluded that dry deposition of sulphate was much greater than wet deposition
at all but one CANSAP sampling site in Alberta (Table 4). In validating the model
Table 2. Wet deposition in Alberta (1978-1984).
Stat i on
Average Annual
Deposition
(mole m"2 y^)
Average
pH
Sul phate
Nitrate
Hydrogen Ion
Beaverlodge
4.7
3.2
5.17
Cal gary
12,8
8.1
0.7
5.77
Coronat i on
7.4
6.3
1 .3
5.46
Edmonton
8.1
6.8
1 .5
5.45
Edson
9,0
4.8
3.1
5.26
Ft. McMurray
e.i
4.8
1 .0
5.61
Lethbridge
10.5
10.8
0.4
6.06
Q Q
O 0 O
1 A
c . u
J . 03
Rocky Mountain House
9.9
6.0
1.5
5.53
Suffield
8.8
6.9
0.8
5.60
Whitecourt
11.4
6.9
1 .1
5.71
Alberta Average
9.2
6.7
1.5
5.48
sulphate: 1 mole wT^ = 0.961 kg ha
nitrate: 1 mole m~^ = 0.620 kg ha
hydrogen ion: 1 mole m ^ = 0.010 kg ha
From: Lau and Das (1985)
9
Table 3. Wet deposition of H+, SOa^", and NOa" (kg ha'^ y^) in Alberta
and at selected Canadian stations from 1978 to 1982.*
Location
Ion
H+
SO42-
NOa-
Beaverlodge (Alta.)
0.031
6.8
3.2
Calgary (Alta.)
0.004
15.5
5.8
Coronation (Alta.)
0.011
8.0
4.6
Edson (Alta.)
0.33
9.0
3.1
Fnr+ MrMiirraw ^Al+a ^
r Kj I L III. 1 lu 1 lay v ' • /
0 010
8 0
0 • u
Lethbridge (Alta.)
0.003
11.5
7.6
Red Deer (Alta.)
0.022
9.4
5.2
Rocky Mtn. House (Alta.)
0.013
10.3
3.7
Suffield (Alta.)
0.009
9.9
4.6
Whitecourt (Alta.)
0.015
10.5
4.3
Prince George (B.C.)
0.009
10.6
3.1
Revelstoke (B.C.)
0.060
6.5
3.9
Cree Lake (Sask.)
0.036
4.4
2.0
Wynyard (Sask.)
0.001
8.2
5.7
The Pas (Man.)
0.003
7.1
3.6
Bissett (Man.)
0.046
7.6
4.1
Moosonee (Ont.)
0.073
11.6
5.6
Simcoe (Ont.)
0.804
52.5
35.0
Maniwalki (Que.)
0.515
32.2
20.3
Quebec City (Que.)
0.573
57.2
25.8
Truro (N.S.)
0.432
30.0
10.9
From Lau (1985) .
10
Table 4. Modelled dry and dry/wet sulphate deposition ratios for Alberta
sites (1982). a
Location
Dry Deposition
(kg ha-i y-i)
Dry-Wet Ratio
Beaverlodge (Alta.)
4.8
0.94
Calgary
20.7
1 .86
Coronation
18.6
2.16
Edmonton
21 .0
1 .94
Edson
18.6
1 .94
Fort McMurray
21 .0
3.18
Lethbridge
13.8
1 .86
Red Deer
12.3
1 .84
Rocky Mtn. House
20.2
1 .94
Suffield
7.2
1 .85
Whitecourt
21 .9
1 .43
Average
16.4
1.86^
3 From Kociuba (1984)
^ The average dry-wet ratio is determined by finding the ratio of the
average dry (16.4 kg ha~^ y~^) and average wet (8.8 kg ha~^ y"^)
depositions .
11
outputs of dry deposition, at this time there is a regrettable lack of sufficient data
on: (a) continuously monitored SO2, and (b) qualitative and quantitative characteris-
tics of fine particulate aerosols. Almost all of the SO2 monitoring in Alberta is
point source oriented in response to regulation.
12
1.5 INTRODUCTION TO ATMOSPHERIC CHEMISTRY AND ACIDIC DEPOSITION PROCESSES:
LITERATURE CITED
Barrie, L.A. 1981. The prediction of rain acidity and SO2 scavenging in eastern North
America. Atmospheric Environment 15: 31-41.
Bielski, B.H.J, D.E. Cabelli, R.L. Arudi, and A.B. Ross. 1985. Reactivity of H202/02~
radicals in aqueous solutions. Journal of Physical Chemistry Reference Data
14: 1041-1100.
Bowersox, V.C. and R.G. de Pena. 1980. Analysis of precipitation chemistry at a central
Pennsylvania site. Journal of Geophysical Research 85: 5614-5620.
Caiazza, R., K.D. Hage, and D. Gallup. 1978. Wet and dry deposition of nutrients in cen-
tral Alberta. Water, Air and Soil Pollution 9: 309-314.
Cass, G.R. and R.H. Shair. 1984. Sulfate accumulation in a sea breeze/land breeze
circulation system. Journal of Geophysical Research 89(D1): 1429-1438.
Chamberlain, A.C. 1986. Deposition of gases and particles on vegetation and soils. In:
Air Pollutants and Their Effects on the Terrestrial Ecosystem, eds. A.H. Legge
and S.V. Krupa. New York: John Wiley and Sons. pp. 189-210.
Chamberlain, J., H. Foley, D. Hammer, G. MacDonald, D. Rothaus, and M. Ruderman. 1981.
The physics and chemistry of acid precipitation. Technical Report JSR-81-25.
SRI International, Palo Alto, California.
Chang, T.Y. 1984. Rain and snow scavenging of HNO3 vapor in the atmosphere. Atmospheric
Environment 18: 191-198.
Chang, T.Y., J.M. Norbeck, and B. Weinstock. 1979. An estimate of the NOx removal rate
in an urban atmosphere. Environmental Science and Technology 13: 1534-1537.
Chapman, E.G., D.S. Sklarew, and J.S. Flickinger. 1986. Organic acids in springtime
Wisconsin precipitation samples. Atmospheric Environment 20: 1717-1725.
Charlson, R.J. and H. Rodhe. 1982. Factors controlling the acidity of natural rainwater.
Nature 295: 683-685.
Cogbill, C.V. and G.E. Likens. 1974. Acid precipitation in the northeastern United
States. Water Resources Research 10: 1133-1137.
Dana, M.T. and R.C. Easter. 1987. Statistical summary and analyses of event precipita-
tion chemistry from the MAP3S network, 1976-1983. Atmospheric Environment 21:
113-127.
Dana, M.T., J.M. Hales, and M.A. Wolf. 1975. Rain scavenging of SO2 and sulphate from
power plant plumes. Journal of Geophysical Research 80: 4119-4129.
Eatough, D.J., R.J. Arthur, N.L. Eautough, M.W. Hill, N.F. Mangelson, B.E. Richter, L.D.
Hansen, and J. A. Cooper. 1984. Rapid conversion of S02(g) to sulphate in
a fog bank. Environmental Science and Technology 18: 855-859.
Finlayson-Pitts , B.J. and J.N. Pitts, Jr. 1986. Atmospheric Chemistry: Fundamentals
and Experimental Techniques. New York: John Wiley and Sons. 1098 pp.
Forrest, J., R.W. Garber, and L. Newman. 1981. Conversion rates in power plant plumes
based on filter pack data: The coal-fired Cumberland plume. Atmospheric
Environment 15: 2273-2282.
Galloway, J.N. and G.E. Likens. 1981. Acid precipitation: The importance of nitric acid.
Atmospheric Environment 15: 1081-1085.
Galloway, J.N., G.E. Likens, and E.S. Edgerton. 1976. Acid precipitation in northeastern
United States: pH and acidity. Science 194: 722-724.
Garland, J. A. 1978. Dry and wet removal of sulphur from the atmosphere. Atmospheric
Environment 12: 349-362.
13
Garrels, R.M. and F.T. MacKenzie. 1971. Evolution of Sedimentary Rocks. New York: Norton.
397 pp.
Gillani, N.V. 1978. Project MISTT: Meso scale plume modelling of the dispersion, trans-
formation, and ground removal of SO2. Atmospheric Environment 12: 569-588.
Graedel, T.E. and K.I. Goldberg. 1983. Kinetic studies of raindrop chemistry. I. Inor-
ganic and organic processes. Journal of Geophysical Research 88: 865-882.
Graedel, T.E., M.L. Mandich, and C.J. Weschler. 1986. Kinetic model studies of atmos-
pheric droplet chemistry. II. Homogeneous transition metal chemistry in
raindrops. Journal of Geophysical Research 91(D4): 5021-5205.
Guiang, S.F., S.V. Krupa, and G.C. Pratt. 1984. Measurement of S(IV) and organic anions
in Minnesota rain. Atmospheric Environment 18: 1677-1682.
Hansen, D.A., G.M. Hidy, and G.J. Stensland. 1981. Examination of the basis for trend
interpretation of historical rain chemistry in the eastern United States.
Report to the Utilities Air Regulatory Group, Committee on Acid Precipitation,
pp. 1-1 to A-1 .
Hegg, D.A. and P.V. Hobbs. 1979. Some observations of particulate nitrate concentrations
in coal-fired power plant plumes. Atmospheric Environment 13: 1715-1716.
Hidy, G.M. and P.K. Mueller. 1986. The sulfur oxide - particulate matter complex. In:
Air Pollutants and Their Effects on the Terrestrial Ecosystem, eds. A.H. Legge
and S.V. Krupa. New York: John Wiley and Sons. pp. 51-104.
Husar, R.B., J. P. Lodge, Jr., and D.J. Moore (eds.). 1978. Sulphur in the Atmosphere.
New York: Pergamon Press. 816 pp.
Hutchinson, T.C. and M. Havas, eds. 1980. Effects of Acid Precipitation on Terrestrial
Ecosystems. New York: Plenum Press. 654 pp.
International Electric Research Exchange. 1981. Effects of SO2 and its Derivatives on
Health and Ecology. 2. Natural Ecosystems, Agriculture, Forestry and Fisheries.
Central Electricity Generating Board, Leatherhead, England, pp. 1-1 to 6-13.
Kasina, S. 1980. On precipitation acidity in southern Poland. Atmospheric Environment
14: 1217-1221.
Kawamura, K. and I.R. Kaplan. 1983. Organic compounds in the rainwater of Los Angeles.
Environmental Science and Technology 17: 497-501.
Keene, W.C. and J.N. Galloway. 1984. Organic acidity in the precipitation of North
America. Atmospheric Environment 18: 2491-2497.
Klemm, R.F. and J. M.L. Gray. 1982. Acidity and chemical composition of precipitation in
central Alberta, 1977-1978. In: Proc. Acid Forming Emissions in Alberta and
Their Ecological Effects. Research Management Division, Alberta Environment,
Edmonton, Alberta, pp. 153-179.
Knapp, W.W., V.C. Bowersox, B.I. Chevone, S.V. Krupa, J. A. Lynch, and W.W. McFee. 1987.
Precipitation chemistry in the United States: Summary of ion concentration
variability, 1979-1984. Progress Reports USDA/Cooperati ve State Research
Service, Washington, DC pp. 1-1 to D-1 .
Kociuba, P.J. 1984. Estimate of sulphate deposition in precipitation for Alberta.
Atmospheric Environment Service, Environment Canada, Edmonton, Alberta. 6 pp.
Krupa, S.V., V.C. Bowersox, B.I. Chevone, W.W. Knapp, J. A. Lynch, and W.W. McFee. 1987a.
Analysis of Inter-regional Project IR-7 Precipitation Chemistry Data. Precipi-
tation Chemistry in the United States. Summary of Qualitative Characteristics.
1979-1984. Progress Report Submitted to USDA-Cooperati ve State Research
Service, Washington, DC. 211 pp.
14
Krupa, S.V., J. P. Lodge, Jr., M. Nosal, and G.E. McVehil. 1987b. Characteristics of aero-
sols and rain chemistry in north central U.S.A. Proceedings of International
Conference on Acidic Rain: Scientific and Technical Advances, Lisbon, Portugal,
eds., R. Perry, R.M. Harrison, J.N.B. Bell, and J.N. Lester. Selper Ltd.:
London, pp. 121-128.
Krupa, S.V., M.R. Coscio, Jr., and F.A. Wood. 1976. Evidence for a multiple hydrogen-
ion donor system in rain. Water, Air, and Soil Pollution 681-683.
Larson, T.V., R.J. Charlson, E.J. Knudson, G.D. Christian, and H. Harrison. 1975. The
influence of a single sulphur dioxide point source on the rain chemistry of a
single storm in the Puget Sound region. Water, Air, and Soil Pollution 4:
319-328.
Lau, Y.K. 1985. A 5-year (1978-1982) summary of precipitation chemistry measurements in
Alberta. Pollution Control Division, Alberta Environment, Edmonton, Alberta.
Lau, Y.K. and N.C. Das. 1985. Precipitation quality monitoring in Alberta. In: Impact
of Air Toxics on the Quality of Life. Proc. 1985 Annual Meeting, CPANS/PNWIS
Sections, Air Pollution Control Association. Pittsburgh, Pennsylvania. pp.
213-233.
Lefohn, A.S. and S.V. Krupa. 1987. The relationship between hydrogen and sulfate ions
in precipitation - A numerical analysis of rain and snowfall chemistry.
Environmental Pollution (In Press).
Lefohn, A.S. and S.V. Krupa. 1984. Sulfate as a surrogate for hydrogen in rainfall - An
analysis of rainfall chemistry data. Draft Report U.S. EPA, Washington, DC.
60 pp.
Legge, A.H. and S.V. Krupa, eds. 1986. Air Pollutants and Their Effects on the Terres-
trial Ecosystem. New York: John Wiley and Sons. 662 pp.
Likens, G.E. and F.H. Bormann. 1974. Acid rain: A serious environmental problem. Science
184: 1176-1179.
Likens, G.E. and T.J. Butler. 1981. Recent acidification of precipitation in North
America. Atmospheric Environment 15: 1103-1109.
Likens, G.E, F.H. Bormann, and J.S. Eaton. 1980. Variations in precipitation and stream
water chemistry at the Hubbard Brook Experimental Forest during 1964-1977.
In : Effects of Acid Precipitation on Terrestrial Ecosystems, Proceedings of
the NATO Conference on Effects of Acid Precipitation on Vegetation and Soils,
eds. T.C. Hutchinson, and M. Havas. 1978 May 21-27; Toronto, Canada; New York:
Plenum Press, pp. 443-464.
Madsen, B.C. 1981. Acid rain at Kennedy Space Center, Florida: Recent observations.
Atmospheric Environment 15: 853-862.
MAP3S/RAINE Research Community. 1982. The MAP3S/RAINE precipitation chemistry network:
Statistical overview for the period 1976-1980. Atmospheric Environment 16:
1603-1631 .
Martin, A. and F.R. Barber. 1977. Some observations on acidity and sulphur in rainwater
from rural sites in central England and Wales. Atmospheric Environment 12:
1481-1487.
McNaughton, D.J. 1981. Relationships between sulfate and nitrate ion concentrations and
rainfall pH for use in modelling applications. Atmospheric Environment 15:
1075-1079.
Meurrens, A. 1974. Sulfites et sulfates dan la pluie. Paper presented at the COST
Technical Symposium, Ispra. (As cited by J. A. Garland, 1978).
Meyers, P. A. and R.A. Hites. 1982. Extractable organic compounds in Midwest rain and
snow. Atmospheric Environment 16: 2169-2175.
15
Miles, L.J. and K.H. Yost. 1982. Quality analysis of USGS precipitation chemistry data
for New York. Atmospheric Environment 16: 2889-2898.
Nosal, M. and S.V. Krupa. 1986. Numerical methodology in the risk assessment of air
pollutant-induced ecological effects. Proceedings of the Annual Meeting of the
Air Pollution Control Association, Pittsburgh, Pennsylvania 86-32.1:1-16.
Nyborg, M., J. Crepin, D. Hocking, and J. Baker. 1977. Effect of sulphur dioxide on
precipitation and on the sulphur content and acidity of soils in Alberta,
Canada. Water, Air and Soil Pollution 7: 439-448.
Peters, N.E., R. Schroeder, and D. Troutman. 1982. Temporal trends in the acidity of
precipitation and surface waters of New York. U.S. Geological Survey Water
Supply Paper 2188.
Pratt, G.C., and S.V. Krupa. 1985. Aerosol chemistry in Minnesota and Wisconsin and its
relation to rainfall chemistry. Atmospheric Environment 19: 962-971.
Pratt, G.C. and S.V. Krupa. 1983. Seasonal trends in precipitation chemistry. Atmos-
pheric Environment 17: 1845-1847.
Pratt, G.C, M.R. Coscio, and S.V. Krupa. 1984. Regional rainfal Ichemistry in Minnesota
and West Central Wisconsin. Atmospheric Environment 18: 173-182.
Pratt, G.C, M.R. Coscio, D.W. Gardner, B.I. Chevone, and S.V. Krupa. 1983. An analysis
of the chemical properties of rain in Minnesota. Atmospheric Environment 17:
347-355.
Richards, L.W., J. A. Anderson, D.L. Blumenthal, A. A. Brandt, J. A. MacDonald, N. Watus,
E.S. Macias, and P.S. Bhardwaja. 1981. The chemistry, aerosol physics, and
optical properties of a western coal-fired power plant plume. Atmospheric
Environment 15: 2111-2134.
Roedel, W. 1979. Measurements of sulfuric acid saturation vapor pressure: Implications
for aerosol formation by heteromolecular nucleation. Journal of Aerosol
Science 10: 375-386.
Schertz, T.L. and R.M. Hirsch. 1985. Trend Analysis of Weekly Acid Rain Data - 1978-1983.
Water Resources Investigations Report 85-4211. United States Geological
Survey. 64 pp.
Sequeira, R.A. 1982. Acid rain: An Assessment based on acid-base considerations. Journal
of the Air Pollution Control Association 32: 241-245.
Snedecor, G.W. and W.G. Cochran. 1978. Statistical Methods. Ames, Iowa: Iowa State
University Press, 593 pp.
Spicer, CW. 1982a. The distribution of oxidized nitrogen in urban air. Science of the
Total Environment 24: 183-192.
Spicer, CW. 1982b. Nitrogen oxide reactions in the urban plume of Boston. Science 215:
1095-1097.
Stensland, G.J., D.M. Whelpdale, and G. Oehlert. 1986. Precipitation chemistry. In:
Acid Deposition. Long Term Trends. Washington, DC: National Academy Press,
pp. 128-199.
Teasley, J.I., ed. 1984. Acid Precipitation Series. Vol. 1-9. Boston Massachusetts:
Butterworth Scientific Publishers.
United States Environmental Protection Agency. 1983. The acidic deposition phenomenon
and its effects. EPA 600/8-83-01 68 . pp. 1-1 to 7-49.
U.S. National Research Council. 1986. Acid Deposition. Long Term Trends. Washington,
DC: National Academy Press. 506 pp.
U.S. National Research Council. 1983. Acidic deposition. Atmospheric processes in east-
ern North America. National Academy Press, Washington, DC 375. pp.
16
Voldner, E.C., L.A. Barrie, and A. Sirois. 1986. A literature review of dry deposition
of oxides of sulphur and nitrogen with emphasis on long-range transport
modelling in North America. Atmospheric Environment 20: 2101-2123.
Whitby, K.T. 1978. The physical characteristics of sulfur aerosols. Atmospheric Environ-
ment 1 2: 1 35-1 59.
Whitfield, M. 1979. Activity coefficients in natural waters. Iri: Activity Coefficients
in Electrolyte Solutions, ed. R.M. Pytkowicz. Boca Raton, Florida: CRC Press,
II: 153-299.
17
2. EFFECTS OF ACIDIC DEPOSITION ON FORESTS
2.1 INTRODUCTION
Sulphur and nitrogen, the two most important components with respect to acidic
deposition, are essential nutrients for plant growth. Sulphur deficiencies are wide-
spread in regions such as the US Great Plains and adjacent Canadian Prairies (Brady
1984). Nitrogen deficiencies have also been postulated for forest soils as a result of
leaching and uptake losses exceeding fixation (Smith 1981). Therefore, both elements
combine fertilizing properties and acid forming capabilities and as a result, have been
the focus of a great deal of research.
Carbon, principally in the form of carbon dioxide, is also essential for plant
growth since it stimulates photosynthetic activity. Atmospheric concentrations of carbon
dioxide have been increasing as a result of fossil fuel combustion, and other industrial
sources, at a rate of about 3% per decade (Smith 1981). Between 1958 and 1978, the
global concentration of carbon dioxide increased from 315 to 335 ppm as a direct result
of anthropogenic activities (Oeschger et al. 1980). It has been estimated that in
pre-industrial times the concentration may have been as low as 265 ppm (Bolin 1983).
These trends have led to concerns over resultant atmospheric warming or the "greenhouse
effect" with potential negative consequences for the biosphere (Bach et al . 1980).
Atmospheric carbon dioxide can also form carbonic acid during transformation
processes. Because it is a weak acid, carbonic acid represents potential acidity and
can act as a buffer. This is in direct contrast to strong acids such as nitric and
sulphuric which dissociate completely (Krug and Frink 1983).
Carbon, like sulphur and nitrogen, is a major stimulator of plant growth which
can have both beneficial and detrimental effects on overall forest productivity. All
three substances can be derived from acidic deposition.
A fourth substance, ozone, is also very important. While it is not not acidic
or an acidifying compound, because of its known direct deleterious effects on plants,
and its joint action with other pollutants including acidic deposition, it must be
considered in the present context. Ozone has been found to have the following effects
on forests:
1. It has been found to cause damage and reduced growth of Ponderosa pine in
the San Bernardino Mountains of southern California (Coyne and Bingham
1977) and has been implicated in white pine emergence tip burn in West
Virginia and Tennessee (Berry and Ripperton 1963; Berry and Hepting 1964).
2. In California, a direct relationship between carbon dioxide increases and
increases in ozone concentration has been found (Coyne and Bingham 1977).
3. It has been speculated that increasing global concentrations of ozone may
also have an effect on temperature similar to that predicted for carbon
dioxide. For example, if the global concentration of ozone were to double,
temperatures could rise by 1°C (Lacis et al. 1981).
4. Ozone has been shown to produce more than additive effects on vegetation
in the presence of other air pollutants, including those associated with
acid emissions (Ormrod 1982).
18
Examples of ozone concentrations reported in the literature are in the range of
666-784 vg for California, 255-911 yg m ^ for Ohio, and as high as 788 yg m ^ in the
Black Forest of West Germany (McLaughlin 1985). Ozone has been implicated as a possible
contributor to forest decline (see definition below) in both Europe and the eastern
United States (Blank 1985).
2.2 FOREST CONCERNS RELATED TO ACIDIC DEPOSITION
Acidic deposition, whether wet or dry, has been documented as occurring in many
areas of the world (McLaughlin 1985). For the purposes of definition, acidic precipita-
tion is any wet event with a pH lower than 5.6, which is the pH of atmospheric waters in
equilibrium with carbon dioxide (Krug and Frink 1983). Postel (1984b) has reported that
in many heavily industrialized areas, precipitation may have acidity levels ten to thirty
times greater than those which would be expected from an atmosphere free of pollution.
However, because of the presence of undi ssociated acids in such precipitation, these
values undoubtedly represent an overestimation of true acidity (Krug and Frink 1983).
In spite of this, acidic rainfall (pH range 4 to 4.6) does occur over much of Europe and
North America (Postel 1984b; Brady 1984).
The possible and potential effects of acidic deposition on forests are discussed
in the following sections.
2.2.1 Forest Decline Phenomenon
There are three categories of plant diseases as defined by Manion (1981):
biotic, abiotic, and decline. Of these disease types, decline is the hardest to define
clearly. Manion, in an attempt at its definition, stated that declines "...result not
from a single casual agent but from an interacting set of factors". McLaughlin (1985)
further stated that declines are complex diseases resulting in a progressive weakening
of trees leading to dieback, and the death of portions of the foliar canopy. Gradual
loss of vigour involving a reduced growth rate and increased susceptibility to secondary
biotic and abiotic stress typically ensues. Declines generally affect mature trees and
ultimately death may occur unless the stress is removed or its effect controlled.
Forests with tree decline typically exhibit a wide and random variety of
symptoms of biotic and abiotic stress (Woodman 1987). Because of the random symptom
distribution patterns displayed in forests undergoing decline, it is almost impossible
to define its exact causes (Hyink and Zedacker 1987).
Many forests in North America and Europe appear to be undergoing change that
includes reduced productivity, dieback, and death. Reviews of these phenomena have been
provided by Abrahamsen et al. (1976), Binns (1984), Morrison (1984), Postel (1984a, b,c,
d.e.f, 1985), and McLaughlin (1985). Stand dynamics and the evaluation of the meaning
of forest decline have been reviewed by Hyink and Zedacker (1987).
Acidic deposition has been implicated as a possible cause for the decline of
forests in Norway (Overrein et al . 1980) and central Europe (Paces 1985; Van Breeman
1985). The decline of West German forests has been documented by Postel (1984a) who
stated that 76% of the fir, 41% of the spruce, and 43% of the pine in the country were
showing symptoms of stress. In total, the apparently stressed timber amounts to
approximately 66% of West German forests. Similar results were documented by McLaughlin
19
(1985) and by Binns and Redfern (1983) in other locations throughout the world. In each
of these cases, acidic deposition was suggested as the most probable cause of forest
decline at the scales noted. In the United States, while a substantial lowering of
forest productivity was very evident, no consensus or compelling documentation as to its
causes were evident (refer to Linthurst 1984). A summary of possible effects of acidic
deposition on soils, plants, and forests is provided in Table 5 along with the references
relevant to these studies.
2.3 DIRECT EFFECTS OF ACIDIC DEPOSITION ON FORESTS
Abundant documentation exists, as shown in Tables 6 and 7, that acid precursors
such as SO2 can cause stomatal closure, affect water relations, reduce photosynthesis,
and change carbon allocation. These effects can be mediated through modifications of
phosphorylation, chlorophyll concentration, carboxylation, hormonal balance, and membrane
integrity. Frank (1985), and Frank and Frank (1985), have also demonstrated that tree
foliar pigments can be destroyed, producing symptoms very similar to those of the German
forest decline, by exposure to chloroethanes . Chloroethanes and other halogenated
chlorocarbons are also associated with acid emission sources and thus, are considered
part of the acidic deposition problem.
Under some conditions, low level exposure of trees to sulphur and nitrogen
oxides can, however, stimulate growth by fertilization. Similarly, many of these acid
precursor pollutants can interact with elevated carbon dioxide, producing either a less
deleterious or even stimulatory effect on forest vigour. As Legge et al. (1986) have
pointed out, the supposed direct effects of the deposition of acidifying pollutants on
biochemical processes in trees as shown in Tables 8 and 9 may, in fact, be indirect
effects caused by the acidification of soils. If this is the case, one can understand
the difficulty for silviculturalists in documenting either the definitive cause of
symptoms or the dose response of vegetation to dry and wet acidic deposition.
2.4 INDIRECT EFFECTS OF ACIDIC DEPOSITION ON FORESTS
There are many indirect effects on forests attributed to acidic deposition.
These include: canopy-pollutant interactions, soil acidification, nutrient leaching,
weathering, and effects on microbial activity. Only canopy-pollutant interactions will
be discussed here since the other topics are discussed at length in other parts of this
synthesis .
2.4.1 Canopy-Pollutant Interactions
Forest canopies act as interceptors of both wet and dry deposition and in the
process, alter the incident chemistry prior to contact with soils. Smith (1981)
estimated that a one hectare model forest was capable of removing the following amounts
of pollutants (t y"^ : Oa - 9.6 x 10*; SO2 - 748; CO - 2.2; NOx - 0.38; and PAN - 0.17.
Binns (1984) reported that up to 40% of precipitation intercepted by the forest canopy
can re-evaporate along with its captured pollutants. This means that these materials
may not pass through to other components of the ecosystem, at least at that location.
Studies of deposition in Sweden on conifers by Granat (1983) showed that sulphur
20
Table 5. References to various effects of acidic deposition on soils,
plants, forests, and ecosystems.
Effects as:
Direct or
Indi rect
Effect
Reference
Direct Effects
Stomatal or mesophyll
resistance
Photosynthesis
Metabolism
Hormones
Membranes
Growth
Black (1982)
Carlson and Bazzaz (1982)
Heath (1980)
Reid (1985)
Skarby and Sellden (1984)
Higginbotham et al. (1985)
Indirect Effects
Canopy leaching
-nutrient leaching
-aluminum and manganese
-phosphorus
— u/p A + h P PI nn
-decomposition
— m\/r n K*i^h 1 7ap
my L (J 1 1 II 1 cav:
-nitrogen fixation/
IIILT MlLaLIUIl
Foster and Morrison (1976)
Morrison (1983)
Ulrich et al. (1980)
Cook (1983)
1nhn<;nn Pt al MQflPh^
Coleman (1983)
P;^ttpn MQR'^^
rauucii ^ ijOO)
Fertilizer effects:
-sulphur
-ammonium
Smith (1981)
Nihlgard (1985)
Harvest technique
Johnson and Richter (1983)
Forest reproduction
Cox (1983)
Land use change and/
or succession
Krug and Frink (1983)
21
22
E o oj
=3 Q. U-
E V4-
•1-0)0)
vi3 CD
CD
cr> r—
O)
O CJ>
C7>
(O O)
c -o 0;
o u
tn O) C
O) CD
O) +->
^ o
o <u o> o
O S- i- o
O) u-
■»-> »4-
O)
s- o
o
O E
C\J Q.
.— CL
O CM
o •
O
oj E
y- Z3
4- E
O) -1-
c
O -1-
z E
O) ro
C -1-7 00
O) -O (O
(J3 ro C71
TO
■M
M
O
C_>
■o
c
■o
o
o;
(_>
ro
(0
>
■a
0)
o>
il
i_
03
o;
+->
I/O
O)
■H
(O
c/)
ro
o>
E
-o
i/i
c
O)
n3
■!->
<_)
C
c
t/1
o
fO
E
■M
x:
1/1
+->
in
c
1/1
c
(O
c=
a;
o
i_
3
03
o
+-)
E
S-
1/1
(O
<
cn
(O
cn
(/)
i-
E
C_)
0)
c
-)->
E
s_
=J
■o
O)
OJ
£=
0)
■M
o
0)
a;
y-
a
(/I
ro
>
c
■M
ro
1/1
'o
c
+J
O)
Q-
(L)
m
=J
E
1/1
ro
rO
s_
O)
>
J-
0)
S_
Q.
+->
X
0)
O
(O
OJ
>
i.
^
l/l
c
■H
o
E*
3
o
j_
o
■o
■o
■H
o
Q
E
OJ
■o
O
s_
>
O)
c
O)
c
O)
4-)
+->
(O
CD
■M
C
<u
<U
m
o
q:
c£:
<_>
_J
H
<N
00
23
c
•r- D.
+-> (V
Q; O
(T3 C
to i-
O O)
+-' C U
O QJ i-
CD 3
CL O O
OJ +-> ro
•r- >, OJ
+
<U ra
cn ^
i- to
It] ^ C CU
O) i- OJ I—
i- O ■(-> -o
CD
■o
■o -o -o
0)0)0)
C)^ Q£
E E E
CL Q. Q.
CL Q. Q.
LTl O LD
t— (D '^f
O O O
+-> O
3 o
0 E s-
S- Q.
01 Q. 1/1
(tJ
E O
=> O
E O r- +J
•r- r- 4-> 3
X l/l O
n3 -H o s_
I CN
I E I
E -o E
-a -o
D)
a,E O)
n
• c\j ■
v£) r- 00
o o o
o o o
n o o
c
o — «.
00 CsJ
I— CD
V- CD
TO r—
■O ^ O
O; r— O
1/1 fO CO
■O ■!-> O
(N O -M E
O S- Z3 Q.
I/) Q.
I— 1/1 -O
l/l Q- O)
i_
<N -O -O -
o <u c:
E -
o.
CL
o o
O • <N
vi3 O O
24
QJ CVJ
C 00
>>cn
o r-
I/) y- -l-i +->
l/l QJ 4- i/i C
0( O <_> CD i-
O •.- C •<- 3
OJ H H
C -H I I
1/1 CT5 C7>
O) cn o o
S- T- C_3 o
O) • ► on I — CM
I — (yn 1/1 ( —
-O Q- to II II
CD >— -H
OJ -M U +-> i>n
cr on on O)
>> 3 on 3 C7>
O di O -t-
OJ Q. on O)
on I Qj C
>> +-> i-H S_
S- ^ +->
Z3 CD - C
•r-)T- O) (U
cr .— -M o
•I- CO (O c
! S- O S-
ro I— I a> c_j u-
CD C -H
E 00
=5 O
r- CD a;
•1— on
E <- I
O on 1—1
i- i- +->
•I— <l) •!— on
CiX: on OJ
on CD C <U
<U -1- di i.
q; ^ on +->
on on o O >
on on O
<L> <D O) M- i- +J
r— > .— O -1-
•r— o3 on
x: +J •» +-> C
+J •!- on i_ on <D on
3 on O) O <U on OJ
O C QJ Q.<— OJ
i- OJ i- X O C i-
CD on +-> <1> ^ •!—
on E * CD
i.
O • +-> u-
^ C u-
>| r^ (O
-M on I +-> -o
•1- <U J= on
(_) <U •!— O
ro i- H on c
Q. -M I O)
(t3 CD S- -
(_> O)
> IN -o cu
on -r- O C •
Q. -M C_J (T3 -1-
£1
y- C E on
on O O Q. I on M_
O) •!— Q. a> OJ o
U C +J S- r—
C O 03 C\J C <N
^ +-> -I- O -r- -r- </0
C •!- Q. fM C_> +->
a> XI on O on o >)
•r- C (/) -H 3 X3
<X. <0 CD r— -o
QDC:S->>-.-=3Qjon
<:t-+->X1^ on S_Q.
E >
Q.
Ql
XJ '3-
<D • C
S_ ,- O
-C) on CL
C CU •!— on
O CJ on
•r- 3 CD ^
-M TD ^ -M
n3 OJ -H •<-
S- 5- c 3
C on on on
CD CD O CD
(_i on -t-J -1-
C =1 O S_
O nD ^ TO
O <_) Q- >
E on OJ
X on CD
TO CD •!- O
3 Q. 00
O
S- >i O
CD -(->
CNJ
0; -r- on +J 4->
CD -O OJ H c
TO I — <D I TO
•I— TO on CD I —
O) O Q.
O +-> <U 1— CD
y- TO c o E E
o
+J H -OH ^ S_
C ! ,— 1 . y-
CM
i- H >, H T3 CD
3 ! I C ^
CJ CD CM CD TO +->
CD O 00
OO on ^
on I-—,
Oi TO
•a
a> I—
«- TO
•1- QJ
<D i/i O
to Q. r—
(T3 C ^
X OJ fO U
O S- -C -f-
2 cn +-> S-
c
+ on "O CU
<D r—
N > OJ <N
O -r- -r- O
C_) CD >> CJ
-l-J -M •» >»
C r— .1- C T3 l/l
O 3 a O r— O
•1- 1/1 i_ •!- O) +->
+-> o) y- +-> o
r— I— Q.
o e o o ^ 3
1/1 Q. on n i- O
Q. -O r— 1—
+-> r— C I +->
c CO QJ -I- ro j::
CD "d- -r- C CD +->
•I- >> 2 -r- S_ -1-
5- C CD 3
■t-> QJ S« E l/l
3 CP n Q. +-> i/l i/l
CO Q-r— 4- -f-
i- + 3 O) 1/1
3 -M O IJO y- QJ
O -r- C r- OJ C ^
_l C •!- I— S_ -1- -M
CU T3 O
C71 OJ >>
O O +
i- =5
+-> -O <N -M
•I- <U O O
-M O C
3 00
O ■(->
I. >1 3
O X) ^
>> fO
>i
TD Xi
OJ Q.
1/1
O OJ
r— Ln
Ln CO
1/1 •(—
>> s-
r- (O
1/1 -C3
to C
CD na
•I- ro
E
C71 <t
a.1—
to XD
•I- 1/1 O (O
-a +->•<- a.
=> c +J E
r— (O <D o
o I— x: o
c Q. +->
•I- c +->
« >> o
« O 1/1 c
Q. ■)->
c:
CJi (U <D
>i > 4- E
O) 1/1
I— OJ
ro E
O) •■-
o +->
(4- O)
CD E
i- o
0) c
» cn
O) =!
cu o
1-
3 •
l/l E
to 3
OJ -1-
E -o
o
(u a
E o
o c
l/l O)
x:
O) <_) _
+-> C fO
(O oj (O 3
o -o s-
•I- a XD c
-D >— O
C O +-> Q.
•1- C 3 3
•I- o
O TD
+-> i>o E QJ
+-> O 1/1
-o C 1- ra
QJ fO y- X3
3 Q. <a to
26
o 3
•I- o
fO CD
J3 -o
—I QJ
OJ 1—
> — '
I c
O -i-
-M
O O
^ QJ
Q. -O
U- Q.
O) CD
1- ^
O O)
y- M Q.
O O -M T3 KJ) ZS h-
t/0 O >— C -t-J <C
C 3 QJ I—
O CnXJ -r- O 3 C71
•I- C <U LL. un <_) C
+-> -r- i_ ^
O IJO X3 O
-O 0)0^ 03 _l ^
CU S- 00 Q.^ r-
<D
I- •!- X Lfl Ol I CT
fO O CM (O Dl fO
<U X: I- CM -r- T-
C 4-> Q. I— CL .—
•r- -r- Q. +J O (— O
— I 3 »a (a u- <t
o r~
un o
CM CM
>> I—
cn
3 JD +->
■O E 5-
OJ OJ ro
S- E CL
1/1 +J i/i +J
fO 03
l/l •!— <_) i- <J i-
a> +-> O) o c o
i/i ra TD JZ ^
fO f— I CL Q.
OJ >5 Ol on +J on
4- S- O O 03 O
O O Z x: <D JC
<u x: CL i- Q.
-o Q. + o cn o
E c I
•I- O (N
o
c
a>
cn
O Ol
s- o
T3 LO
>y
on -H 2 -t-J Ol 1-
C CD 1/1 C
Ol T3
(/> O Q)
■H
It) ^ O)
<J C +->
•1- O •«-
u a> la
C l/l O)
•r- n3 1-
-)-> o
-(-> o c
CO Z3 -r-
1- QJ O
Ol 4- C
to (TJ
I— OJ
"O QJ
C i- -o
<a c
on fO
on QJ
S- -O on
ro •!— Q-
cn i- Q.
=J TCI -O O
on ^ QJ t.
u u TD
T3 U 3
QJ (O -O O)
(_) on QJ 1 —
3 >i i- -O
"Or— QJ
QJ O O QJ
cn CL -H C
(a fO >)
CD on ^
3 >i
un 1 — "D
QJ O <_)
QJ O)
c >i-a
cn fo "o fO
E
CL on
Q. >i
no
O T3
cn
• o
o m
27
a;
J3 TD
r—
—I O)
3 C OJ
S_ (O i_
>» -H
+-> 1/1 ro 1/1
•r- I -(-> OJ
i- fo «a i/»
cj) s- e i>o 1/1
<u +-> O OJ -1-
-♦-> I— ■(-> O 1/1
C 3 1/1 O O)
•I- i- ^
(D C Q. +->
O) >— O C
C r— 1/1 >,
(O QJ ■)-> =J 1/1
i- C U O O
^ H3 <U T- +J
E cny- i- o
Q> 4- 14- to ^
S O (U > Q.
I 3 ^ .— ■♦-> J3 -1- "O
•I- o -I- >) e q; •/!•<- la
«t <_) (O Q. "O
CD 1- O C
< •— s- »a
(TJ C O
i/i -l-J o
r- I/O QJ J3 1— Q.
x: ■»-> 3 i- O) >—
o i/i 1/1 ro > 3
<U Q. T- Q. -r- ♦ ^
■O O 1. UJ +J (NO.
to S_ -O Q- Q) O I —
S- O C Q. 00 3
Oil— o -o E 1/1
(U ^ ^ C O >>•!-
■a o o ca <_> J2
■O I r- t—
•1-1/1 • l/l
3 0, ► a>
l<0 c o «
(t) c: o c: (u
O -r- O) C
l/l Q. l/l O
1/1 O) >
3 i- -1-
<a u
O r-
OJ I—
C >> O)
O) +-> O
r— O)
>,•<- 3
x: i- o
-M fO r-
UJ > l/l
C (O
i- O Q.
O) •>- X
3 -M O)
O to
.- E «+-
14- S_ ro
oO O Q.
O) Q.
H H C E "O
II (U -O Q..—
_J _l r— -r- Q. O
>i O ^
I I ^ ro O l/l
3. 3. +-> o a>
O) CM i-
tn CO c JZ
o • <u o o +->
• O +->•<- +J
O ro -H ^
■H r- <j
4-> lO 3 =3 Q. r—
ro E T3 Q.
n -t- O 1_
10 O -H i- O O
O (/) i/> Q. If) <4-
O. Ji^
E o
>> ro
l/l X3
■O -r-
Q> -O
Q.
0 C
I— ro
01 E
> t.
(U O)
O dJ
■M C
-o x:
Q) +->
l/l a>
o o
Q. i.
X o
QJ 1—
..- ro
Ql O
+J >>
c_3 -I- a -o
— I E -M <u
Q- •!- ro a>
:r l/l c c
S- Q>
+-> c
28
o sz
a. 4- cn
S_ O) -1-
x: <u 3 <D
3 o 3 c
•r- <U r— CU
^ M- -O r—
- <D r—
^ -O CU O) O
+J (O C t/0 Q.
•r- ^ O
t/1 O 1/1 fO S-
<u O) 3 i-
O) > o <u
lyl to I — Dl
fO r— fO r—
OJ I— <D I—
S- O i- O
<s. a. <: Ql
CD C 01 C
> a> > o)
O Ol^ (J 05S«
\ o \ n
(yn C\J v£) 00 O CD
-o CO "o ur>
<u ir> c= <u 00 c=
O) • o a; • o
(/I I •!- 1/1 I —
t-~ +-> 03
o ^
i/i r-
O r—
O i-
Q. i~ - 0)
i- <U +-> CJ)
J= O) 3 ^
D5 3 O D1 C
•I- <U r— -r- <D
^ M_ QJ I—
- 3 r-
x: -o O) o
+J (O C T3 Q.
•(- ^ O O)
3 o cu ^
i/i O i/i i-
ro -r- -O • <U
oj +-> O) CP 3
I- 3 O) > O
<C I — t/) (O 1 —
■H O
3 +^
+-> • S« O J3
050 O in -3
3 -M
a
ra 1—
O) r—
i- o
<: CL
r~ cn
in
o ^ <u
00 ^ X3
) • 00 ^ 00 a
v£) v£) +->
OJ -H • -t-
■— 3 -o c E c CM
O) •<— i_ 0)01
O) TO <u E <u r—
c 0) 1/1 I- Di r— x:
O <U <D O +->
<_> 1/1 -O CD C Q.
(DOC
. . .— -O r— . O)
CD CDi— O) r— cni—
> > -1- OJ o >
<: <t Ll_ 00 Q- <
3 -o
Oi -o O) E O) I—
c <u i- CDi— x:
O OJ O) o +->
CD c Q. cn
<D C
-O I— • O)
CD CD<— Q) "— oil—
> > -r- O) O >
< <C U_ C/1 Q. <C
O) 3 O CD
E O o >
l/l I — TO
i- ► J= -o
CD ^ ■M c:
Q. <N CD (O
Q. O C
O t/1 cu -
O ,— H-J +->
x: jz sz
1- CD ai CD CD
fO •■- C -r- -i-
<U ^ O CLI <D
z >- u 3 3
in in CD iD c\j CD
I CO I "^l-
in I CO r— I 1 —
00 1 — o in CD
«d- CM CO in CD in
■O CD C\J "O CD CO
CD C C CD C C
> o o > o o
<a: c_j c_5 <t cj cj
+-> ^
cr> •!- o
=3 O x: s-
o -o 3-1-
fa <_> •
y- c <L> ofl
O OJ <u r—
(_) Q. -
C C t- fO i-
O Qi <U E "3
•I- D1 E -o
■H i- (D - <U
(O <U >) <_)
.- E x: s-
=3 4-> O »
E 3 cu
i. o <_> c
+J ..- i- .,- .r-
t/) <+- CT) ^ Q.
= <o
C •
•I- tn
i- O
■H
•1- n
E ■»->
=
c •
•r- LD
to
S- O
■M
•1- OO
U
= Ol
C CD C
(O r- TCI
> ^ >
1X5 CD
E >
E
1-
<D CNJ
C\J =
I C
in -t-
-•t- XI (O
-OX) E
o 3
o ^ o s_
3 o r- to
CJ> CO 1 — CD
O I— QJ =3
Q CO >- I/)
C T3 >)
CO 0^ -M
<_) •!—
C 3 >
-O O "O •<-
0)
-!-> ta: Q.
X) c o
•I- <u 3 u
cn o -r-
• c cn n3
CO O) E
Vl>— <V CD
1— X3 •!-
IE o a +->
Q. Q. +-> 1/1
.,- +J 1/1
OJ -M C +->
l/i rc O <D
C I— T- C
O 3 +-> -r-
Q. E •.- E
1/1 -e- X3 i-
<U ■(-> •!- <U
a: i/i
E
s- a> «3-
Ol on ^
E c E
•1- O i-
4-> .r- O)
J- cr^
<:lh-
3 ^
O f~-
r— <yi
30
■»-> o
t/l 3 O)
o -o ■(->
E •-- (O
i- -a <u
<v B
M- - S-
•I- +-> <u
■H 3
•I- o
^ s-
•r- CD
O O
in Q.
O 01 r— vO CM CTl
(O
= n CO CO n CO c\j
+-> O)
c o
o «tl
31
compounds were deposited in wet form in greater amounts than dry. Yearly average depo-
sition rates were 0.90 and 0.32 g for wet and dry, respectively. In contrast, Hofken
(1983) found that dry deposition exceeded wet deposition in a spruce forest and the rates
varied by season. For example, he observed that the rates for dry sulphate deposition
were 1400 and 2600 mg m ^ month ^ for summer and winter, respectively. Corresponding
seasonal rates for wet deposition were 300 and 600 mg m ^ month ^.
Results obtained by Hofken (1983) for nitrogen compounds in wet and dry
deposition were similar to those for sulphur. These data indicate that considerable
amounts of both wet and dry deposition, particularly acidic ions, can be intercepted by
leaves and stems and can undergo reactions with these vegetation parts and later be
washed from the plant to undergo further interaction at the soil interface.
Throughfall may be defined as that portion of precipitation intercepted by the
canopy but subsequently transferred to the soil below. The effects of this interception
vary with the type of forest, i.e., hardwood versus softwood, as well as by species of
tree involved. In general, hardwood canopies tend to raise the pH of the throughfall
(Abrahamsen etal. 1976; Hoffman etal. 1980; Miller 1983; and Mollitor and Raynal
1983), while coniferous or softwood forests generally tend to lower pH. However, Miller
(1983) reported that young Scots pine stands and Sitka spruce both appear to raise,
rather than lower, pH.
Hoffman et al. (1980) reported that total acidity of throughfall was approxi-
mately the same as incident precipitation but that weak acids increased by 20-40% in the
throughfall while strong acids decreased by similar amounts. This finding suggests that
weak acids were exchanged for strong acids.
While hardwood canopies tend to decrease the H^ concentration in the through-
fall (Mollitor and Raynal 1983), total cation concentrations increase (Eaton et al. 1973)
probably as a result of H^ to cation exchange occurring in the canopy through contact
between leaves and the precipitation. Eaton et al. (1973) found in their study that the
concentration of sulphate also increased, thus providing a mobile anion to balance
leaching cations. In comparison, concentrations of both nitrate and ammonium were lower
in the throughfall than in incident precipitation, which suggested differential
absorption (Eaton et al . 1973).
In general, conifers have been found to cause H^ and cation concentrations to
increase in throughfall. Again, this trend can be species specific as shown by Cole and
Johnson (1977) and Binns (1984) in studies with Sitka spruce and Douglas fir respec-
tively. Throughfall from these species is characterized by decreases in h"*" and
increases in the concentration of SO*^", K"*", Ca^"*", and Mg^"*" in comparison with values
found in the precipitation (Eaton et al. 1973). The studies of Miller (1983) on Scots
pine indicated that this effect could also be age specific, with older trees of this
species causing an increase in throughfall H^ , while H^ decreased in the throughfall
of young trees.
Stemflow refers to that portion of the intercepted precipitation that is carried
down the stems, branches, and trunks of trees to the ground. In most cases stemflow in
Scots pine, Sitka spruce, Japanese larch, birch (Miller 1983), lodgepole pine, and white
spruce (Baker et al . 1977) was more acidic than throughfall. In all cases cited, stem-
flow had a lower pH than incident precipitation and decreased even when canopy effects
on throughfall indicated a rise in pH.
32
The possibilities for indirect effects of acidic deposition on forests are
almost limitless. From the moment that acidic materials come in contact with the canopy,
various effects are possible. The canopy itself can exchange and modify precipitation
in a number of ways, so that whatever reaches the ground may be different in its pH,
molecular species of the acids, ionic content, and concentration in comparison with the
incident precipitation. Upon entering the soil, precipitation is subjected to what is,
in effect, a large, variable, and reactive exchange column. The soil can be acidified
with increases in Al^^, Mn^^, or Fe^"*" to toxic levels. Cations may be leached and
microbial processes altered. The whole nitrogen budget may also be altered. Nodule
formation may be inhibited, nitrogen fixation altered, and nitrification inhibited. The
size and proportions of the nitrate-ammonium pools may also change and mycorrhizal
relations may be altered. These possibilities are site-, species-, and situation-specific
and each situation, therefore, is different and requires investigation prior to formu-
lating opinions on possible effects and/or impacts that result from acidic deposition.
There is little doubt that these kinds of effects can occur; what is in doubt is their
importance to overall forest health and under what specific conditions they occur. For
example, although NOx and its family of compounds are often cited as affecting vegetation
as a result of acidic deposition, Vitousek et al. (1979) found that trenching and cutting
alone were sufficient to cause nitrate losses from the soils of 19 different forests.
Presumably, the loss or leaching of the mobile anion nitrate was balanced by an equiva-
lent loss of cations to maintain the system in equilibrium. Thus, disturbance activities
in forest ecosystems tend to mimic the results of acidic deposition.
2.5 INTERACTIVE EFFECTS OF ACIDIC DEPOSITION ON FORESTS
2.5.1 Interactive Effects on Forest Nutrition and Growth
Along with the direct effects of acidic deposition, a wide array of indirect
effects are also possible. Such effects result from forest harvest practices, fertili-
zation, and nutrient cycling, as they interact with acidic deposition.
Abrahamsen (1980) has reviewed the potential relationships between acidic
deposition and plant nutrition. Using Mi kael i s-Menton kinetic theory, he has pointed
out that most forests are limited by the availability of nitrogen and that an over-
abundance of other nutrients cannot compensate for the deficiency of one element or
limitation to growth. Therefore, in such a case, adding excess sulphur will not promote
growth even though it is a plant nutrient. Instead, it may cause additional problems;
fertilization with one element may cause deficiencies in others. For example, the
additions of NH4^ to European forests as a result of acidic deposition may be the
cause of much of the observed forest dieback or decline (Nihlgard 1985). Discussions of
nutrient cycling and supply with respect to forest management practices may be found in
the reviews of Khanna and Ulrich (1984), Miller (1984), and Gosz (1984). In spite of
this body of knowledge, considerable disagreement still exists over the exact causes of
forest decline. For example, Tabatabai (1985) feels that pH changes that result from
acidic deposition are minimized by the natural buffering capacity of soils and that the
resulting additions of N and S will be beneficial. It appears from his discussion that
near pollution sources, deposition may cause adverse effects on crops through the
33
acidifying process, but at long distances from sources, deposition does not appear to be
a concern. Mayo (1987), in his review, documents some of the major investigations that
show beneficial, detrimental, or no effects on vegetation as a result of acidic deposi-
tion. Many of the studies which have shown either beneficial or no effects of acidic
deposition dealt with cereal crops in fumigation experiments carried out in greenhouses
with artificial acidic rain. As Tabatabai (1985) stated in his review, experiments with
simulated acidic rain will not provide the information needed to clearly define effects
on a ambient deposition dose-response basis. Many of the studies that showed detrimental
effects as a result of exposure to gaseous pollutants were conducted in the field and
all dealt with trees of commercial interest. The results of these latter studies clearly
showed that second and third order effects were often found in the trees. It may be
that this degree of interaction was not tested in the studies cited as showing benign
effects of acidic deposition on vegetation. It is also possible that the artificial
"acidic rain" concocted for use in experiments may have led to simplistic and unrealistic
results, in terms of the real world. Such studies are, however, of value as long as
they are then proven under ambient conditions in the field with the same results. An
example of such an experimental approach may be found in the studies of Legge et al.
(1976) with sulphur dioxide, where laboratory fumigation studies were repeated in the
field close to a sulphur gas emitting source and both lab and field experiments produced
similar results.
2.5.2 Timber Harvesting and Acidic Deposition
Forest management practices have been found to affect nutrient cycling (see for
example Vitousek et al. 1979). Variables of concern include the type of harvest, such
as whole tree or bole only, and length of rotation between cuts. For example, an 80-year
rotation with only saw log removal will produce very different results with respect to
nutrient losses than will whole-tree harvest biomass systems of management.
Johnson and Richter (1984) have reviewed the effects of harvesting and pollution
on several forest ecosystems in the United States and West Germany. They reported that
clearcutting operations increased the leaching of both nitrate and calcium. They also
reported that whole tree harvesting results in significantly higher losses of N, S, Ca,
K, and Mg from the system than does bole harvesting. During their studies it was noted
that sulphur input in areas subject to acidic deposition was higher than removal rates
by leaching but that this was not so for nitrogen (Johnson et al. 1982a; Johnson and
Richter 1984). They also concluded that both harvesting and acidic deposition can
result in base cation loss and that, of the two processes, harvesting was likely to be
the most detrimental. These studies reaffirm the conclusions put forth by Tabatabai
(1985) that the nutrient in shortest supply dictates the form of the effects on the
system. In the aforementioned studies, nitrogen was found to be the limiting factor.
Forest management practices would likely be more detrimental if acidic deposition did
not occur because of the losses of nitrogen due to methods such as clear cutting. These
nitrogen losses are to some extent ameliorated by nitrogen deposition in the form of
acidic precipitation and dry deposition.
34
2.5.3 Effects of Acidic Deposition on Tree Reproduction
A summary of the effects of acidic deposition on the reproductive processes of
trees is shown in Table 9 (Mayo 1987). Generally, pollutants reduce pollen germination
and tube growth and there is considerable evidence that cone size, weight, and numbers
of seeds/cone are also reduced (Scheffer and Hedgecock 1955; Cox 1983, 1984; DuBay and
Murdy 1983; and Van Ryne and Jacobson 1984). The results of investigations into effects
on seed germination are contradictory. For example, Bonte (1982) and Raynal et al.
(1982) found that acidic precipitation and high pollutant levels inhibited germination
in red pine and red maple, respectively, while Lee and Webber (1979) and Raynal et al.
(1982) found a stimulatory effect for fir and white pine. Seedling emergence studies
seem to indicate inhibition by acidic precipitation for most species while ozone appears
to stimulate emergence at least for Ponderosa pine (Wilhour and Neely 1977). From the
literature it appears that pollutants such as SO2, ozone, and acidic precipitation can
affect reproduction; however, these effects are quite varied and are distinctly species
and site specific.
2.6 EFFECTS OF ACIDIC DEPOSITION ON PLANT COMMUNITIES
Few studies were reviewed that focused on whole forest effects and responses to
acidic deposition. Very few forest plant community studies have been completed.
However, population studies and ecotype investigations have been completed and the
changes in reproductive biology described in the previous section would suggest that
varied responses leading to community changes may result from acidic deposition. Law
and Mansfield (1982), in studies on NOx, concluded that varietal resistance to nitrogen
uptake may cause variable uptake rates. Studies by Garsed and Rutter (1982) showed wide
differences in sensitivity to SO2 among three different species of pine. Their studies
suggested that pollution could exert a strong selective pressure both within and between
species. Chronic pollution has also been shown to cause simplification of community
structure (McClenahen 1978). All of these studies suggest that chronic low level
pollution could affect forest community diversity and perhaps even dominance structure.
Pielou (1982), however, when testing the species composition along potential pollutant
gradients adjacent to an Oil Sands plant in Alberta, found no evidence of compositional
changes .
It has already been mentioned that there are those who feel that secondary
succession is an acidifying process and that changing land use involving succession back
to coniferous growth may be a major cause of soil acidification (Rosenqvist 1978;
Overrein et al . 1980; and Krug and Frink 1983). Therefore, rather than acidic deposition
altering succession, succession may be the cause of soil acidification in some instances.
35
2.7 EFFECTS OF ACIDIC DEPOSITION ON FORESTS: LITERATURE CITED
Abrahamsen, G. 1980. Acid precipitation, plant nutrients, forest and growth. In:
Ecological Impact of Acid Precipitation, Proceedings of an International
Conference, eds. D. Drablos and A. Tollan. 1980 March 11-14; Sandefjord,
Norway; SNSF Project, Oslo, Norway; pp. 58-63.
Abrahamsen, G., K. Bjor, R. Horntvedt, and B. Tveite. 1976. Effects of acid precipitation
on coniferous forests. In* Impact of Acid Precipitation on Forest and Fresh-
water Ecosystems in Norway. SNSF Research Report No. 6, ed. F.H. Braekke. Oslo,
Norway: SNSF Project, pp. 36-63.
Bach, W., J. Pankrath, and J. Williams, eds. 1980. The Interactions of Energy and
Climate. Boston, Massachusetts: D. Reidel. 569 pp.
Baker, J. 1977. Nutrient levels in rainfall, lodgepole pine foliage and soils surrounding
two sulfur gas extraction plants in Strachan, Alberta. Canadian Forest Service,
Information Report NOR-X-194. Edmonton, Alberta: Northern Forest Research
Centre. 18 pp.
Baker, J., D. Hocking, and M. Nyborg. 1977. Acidity of open and intercepted precipita-
tion in forests and effects on forest soils in Alberta, Canada. Water, Air,
and Soil Pollution 7: 449-460.
Beckerson, D.W. and 6. Hofstra. 1979. Response of leaf diffusive resistance of radish,
cucumber and soybean to Oa and SO2 singly or in combination. Environment
13: 1263-1268.
Belser, L.W. 1979. Population ecology of nitrifying bacteria. Annual Review of Micro-
biology 33: 309-333.
Berry, C.R. and G.H. Hepting. 1964. Injury to eastern white pine by unidentified atmos-
pheric constituents. Forest Science 10: 2-13.
Berry, C.R. and L.A. Ripperton. 1963. Ozone, a possible cause of white pine emergence
tipburn. Phytopathology 53: 552-557.
Biggs, A.R. and D.D. Davis. 1982. Effects of sulfur dioxide on water relations of hybrid
poplar foliage and bark. Canadian Journal of Forest Research 12(3): 612-616.
Binns, W.O. 1984. Acid rain and forestry. Research and Development Paper No. 134.
Edinburgh, Scotland: Forestry Commission. ISBN-0-85538-1 77-9 . 17 pp.
Binns, W.O. and D.B. Redfern. 1983. Acid rain and forest decline in Western Germany.
Research and Development Paper 131. Wrecclesham, England: Forestry Commission
Research Station. 13 pp.
Black, V.J. 1982. Effects of sulphur dioxide on physiological processes in plants. In:
Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterworth Scientific, pp. 67-91.
Blank, L.W. 1985. A new type of forest decline in Germany. Nature 314: 311-314.
Bolin, B. 1983. The carbon cycle in the major biogeochemical cycles and their inter-
actions. I_n: The Major Biogeochemical Ccyles and Their Interactions, eds. B.
Bolin and R.B. Cook. Chichester, UK: John Wiley and Sons, pp. 41-45.
Bonte, J. 1982. Effects of air pollutants on flowering and fruiting. In.* Effects of
Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H. Unsworth and
D.P. Ormrod. London: Butterworth Scientific, pp. 207-223.
Brady, N.C. 1984. The Nature and Properties of Soils. 9th ed. New York: MacMillan.
750 pp.
Bytnerowicz, A. and D.C. Taylor. 1983. Influence of ozone, sulfur dioxide and salinity
on leaf injury, stomatal resistance, growth and chemical composition of bean
plants. Journal of Environmental Quality 12(3): 397-405.
36
Caput, C, Y. Belot, D. Auclair, and N. Decourt. 1978. Pine depressed transpiration by
SO2. Environmental Pollution 16: 3-15.
Carlson, R.W. 1982. The influence of elevated atmsopheric CO2 and SO2 on the growth and
photosynthesis of early successional species. In: Effects of Gaseous Air
Pollution in Agriculture and Horticulture, eds. M.H. Unsworth and D.P. Ormrod.
London: Butterworth Scientific, pp. 489-491.
Carlson, R.W. and F.A. Bazzaz. 1982. Photosynthetic and growth response to fumigation
with SO2 at elevated CO2 for Ca and C4 plants. Oecologia 54: 50-54.
Cole, D.W. and D.W. Johnson. 1977. Atmospheric sulfate additions and cation leaching in
a Douglas fir ecosystem. Water Resources Research 13: 313-317.
Coleman, D.C. 1983. The impacts of acid deposition on soil biota and C cycling. Envi-
ronmental and Experimental Botany 23(3): 225-233.
Cook, R.B. 1983. The impact of acid deposition on the cycles of C, N, P, and S. In.:
The Major Biogeochemical Cycles and Their Interactions, eds. B. Bolin and R.B.
Cook. SCOPE 21. Chichester, UK: John Wiley and Sons, pp. 345-364.
Cox, R.M. 1984. Sensitivity of forest plant reproduction to long range transported air
pollutants: in vitro and in vivo sensitivity of Oenothera parvi flora L. pollen
to simulated acid rain. New Phytologist 97: 63-70.
Cox, R.M. 1983. Sensivity of forest plant reproduction to long range transported air
pollutants: in vitro sensitivity of pollen to simulated acid rain. New
Phytologist 95: 269-276.
Coyne, P.I. and G.E. Bingham. 1982. Variation in photosynthesis and stomatal conductance
in an ozone-stressed ponderosa pine stand: Light response. Forest Science 28:
257-273.
Coyne, P.I. and G.E. Bingham. 1977. Carbon dioxide correlation with oxidant air pollu-
tion in the San Bernardino Mountains of California. Journal of the Air Pollution
Control Association 27(8): 782-783.
DuBay, D.T. and W.H. Murdy. 1983. The impact of sulfur dioxide on plant sexual repro-
duction: in vivo and in vitro effects compared. Journal of Environmental
Quality 12(1): 147-149.
Eaton, J.S., G.E. Likens, and F.H. Bormann. 1973. Throughfall and stemflow chemistry in
a northern hardwood forest. Journal of Ecology 61: 495-508.
Eckert, R.T. and D.B. Houston. 1980. Photosynthesis and needle elongation response of
Pinus strobus clones to low level sulfur dioxide exposures. Canadian Journal
of Forest Research 10: 357-361.
Farrar, J.F., J. Relton, and A.J. Rutter. 1977. Open stomata in pine. Journal of Applied
Ecology 14: 861-875.
Foster, N.W. and I.K. Morrison. 1976. Distribution and cycling of nutrients in a natural
Pinus banksiana ecosystem. Ecology 57: 110-120.
Frank, H. 1985. The new forest decline. Destruction of pigments by UV-acti vation of
chloroethenes : a probable cause of forest decline. US Environmental
Protection Agency Newsletter (April/June): 7-13.
Frank, H. and W. Frank. 1985. Chlorophyll bleaching by atmospheric pollutants and sun-
light. Naturwissenschaften 72: 5139.
Garsed, S.G. and A.J. Rutter. 1982. The relative sensitivities of conifer populations
to SO2 in screening tests with different concentrations of sulphur dioxide.
In: Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterworth Scientific, pp. 474-475.
37
Gosz, J.R. 1984. Biological factors influencing nutrient supply in forest soils. In:
Nutrition of Plantation Forests, eds. G.D. Bowen and E.K.S. Nambiar. London:
Academic, pp. 119-146.
Granat, L. 1983. Measurements of surface resistance during dry deposition of SO2 to wet
and dry coniferous forest. Iji: Effects of Accumulation of Air Pollutants in
Forest Ecosystems, eds. B. Ulrich and J. Pankrath. Boston, Massachusetts: D.
Reidel, pp. 83-89.
Harvey, G.W. and A.H. Legge. 1979. The effect of sulfur dioxide upon the metabolic level
of adenosine triphosphate. Canadian Journal of Botany 57(7): 759-764.
Heath, R.L. 1980. Initial events in injury to plants by air pollutants. Annual Review
of Plant Physiology 31: 395-491.
Higginbotham, K.O., J.M. Mayo, S. L ' Hi rondel le, and D.K. Krystofiak. 1985. Physiological
ecology of lodgepole pine Pinus contorta in an enriched CO2 environment.
Canadian Journal of Forest Research 15: 417-421.
Hoffman, W.A., Jr., S.E. Lindberg, and R.R. Turner. 1980. Precipitation acidity: The
role of the forest canopy in acid exchange. Journal of Environmental Quality
9: 95-100.
Hofken, K.D. 1983. Input of acidifiers and heavy metals to a German forest area due to
wet and dry deposition. In: Effects of Accumulation of Air Pollutants in
Forest Ecosystems, eds. B. Ulrich and J. Pankrath. Boston, Massachusetts: D.
Reidel, pp. 57-64.
Houston, D.B. and I.S. Dochinger. 1977. Effects of ambient air pollution on cone, seed,
and pollen characteristics in eastern white and red pines. Environmental
Pollution 12: 1-5.
Hyink, D.M. and S.M. Zedaker. 1987. Stand dynamics and the evaluation of forest decline.
Tree Physiology 3: 17-26.
Johnson, D.W. and D.D. Richter. 1984. The combined effects of atmospheric deposition,
internal acid production, and harvesting on nutrient gains and losses from
forest ecosystems. In: Forest Responses to Acidic Deposition, Proceedings of
U.S. -Canadian Conference, ed. C.S. Cronan. 1983 August 3-4; Orono, Maine; pp.
36-46.
Johnson, D.W. and D.D. Richter. 1983. The combined effects of atmospheric deposition,
internal acid production, and harvesting on nutrient gains and losses from
forest ecosystems. I_n: Forest Responses to Acidic Deposition, Proceedings of
U.S. -Canadian Conference, ed. C.S. Cronan. 1983 August 3-4; Orono, Maine; 28 pp.
Johnson, D.W., D.C. West, D.E. Todd, and L.K. Mann. 1982a. Effects of sawlog vs. whole-
tree harvesting on the nitrogen, phosphorus, potassium, and calcium budgets of
an upland mixed oak forest. Soil Science Society of America Journal 46:
1304-1309.
Johnson, D.W., J. Turner, and J.M. Kelly. 1982b. The effects of acid rain on forest
nutrient status. Water Resources Research IB: 449-461.
Kelly, J.M., R.C. Strickland, F.P. Weatherford, and J.C. Noggle. 1984. Evaluation of
simulated acid precipitation effects on forest microcosms. Electric Power
Research Institute Report EA-3500. Muscle Shoals, Alabama: Tennessee Valley
Authority. 202 pp.
Khanna, P.H. and B. Ulrich. 1984. Soil characteristics influencing nutrient supply in
forest soils. In: Nutrition of Plantation Forests, eds. G.D. Bowen and E.K.S.
Nambiar. London: Academic, pp. 79-117.
Krug, E.C. and C.R. Frink. 1983. Effects of acid rain on soil and water. Connecticut
Agricultural Experiment Station Bulletin 811, p. 45.
Lacis, A., J. Hansen, P. Lee, T. Mitchell, and S. Lebedoff. 1981. Greenhouse effect of
trace gases. Geophysical Research Letters 8: 1035.
38
Law, R.M. and T.A. Mansfield. 1982. Oxides of nitrogen and the greenhouse atmosphere.
In: Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterowrth Scientific, pp. 93-112.
Lee, J.J. and D.E. Weber. 1979. The effect of simulated acid rain on seedling emergence
and growth of eleven woody species. Forest Science 25: 393-398.
Legge, A.H. and J.C. Bogner. 1983. Ecological monitoring of sulphur in forests in western
Canada. Aquilo Series Botanica 19: 119-139.
Legge, A.H., J. Corbin, J. Bogner, M. Strosher, D. Parkinson, H.R. Krouse, E.J. Laishley,
M.J. Cavey, C.E. Prescott, R.J.F. Bewley, M. Nosal, H.U. Schellhase, T.C.
Weidensaul, and J. Mayo. 1986. Acidification in a temperate forest ecosystem:
The role of sulphur gas emissions and sulphur dust. A final report submitted
to the Whitecourt Environmental Study Group, August 1986. 471 pp.
Legge, A.H., D.R. Jaques, G.W. Harvey, H.R. Krouse, H.M. Brown, E.C. Rhodes, M. Nosal,
H.U. Schellhase, J. Mayo, A. P. Hartgerink, P.P. Lester, R.G. Amundson, and
R.B. Walker. 1981. Sulphur gas emissions in the boreal forest: The West
Whitecourt case study I. Executive Summary. Water, Air, Soil Pollut. 15: 77-85.
Legge, A.H., G.W. Harvey, P.P. Lester, D.R. Jaques, H.R. Krouse, J.M. Mayo, A. P.
Hartgerink, R.G. Amundson, and R.B. Walker. 1976. Quantitative assessment of
the impact of sulfur gas emissions on a forest ecosystem. A final report
submitted to the Whitecourt Environmental Study Group, March 1976. 159 pp.
Linthurst, R.A., ed. 1984. Direct and indirect effects of acidic deposition on vegeta-
tion. Acid Precipitation Series, Vol. 5. Boston, Massachusetts: Butterworth
Scientific. 117 pp.
Majernik, 0. and T.A. Mansfield. 1970. Direct effect of SO2 pollution on the degree of
opening of stomata. Nature 227: 377-378.
Malhotra, S.S. 1977. Effects of aqueous sulfur dioxide on chlorophyll destruction in
Pinus contorta. New Phytologist 78: 101-109.
Manion, P.D. 1981. Tree Disease Concepts. Prentice-Hall, Inc. Englewood Cliffs, New
Jersey. 399 pp.
Mayo, J.M. 1987. The Effects of Acid Deposition on Forests. Prep for the Acid Deposition
Research Program by the Department of Biology, Emporia State University,
Emporia, Kansas. ADRP-B-09-87 . 80 pp.
Mayo, J.M., A. P. Hargerink, and A.H. Legge. 1986. Sulphur gas emissions in the boreal
forest: The West Whitecourt case study VI: Woody plant water status. Water,
Air, and Soil Pollution 29(2): 113-127.
McClenahen, J.R. 1978. Community changes in a deciduous forest exposed to air pollution.
Canadian Journal of Forest Research 8: 432-438.
McLaughlin, S.B. 1985. Effects of air pollution on forests: A critical review. Journal
of the Air Pollution Control Association 35(5): 512-534.
McLaughlin, S.B., R.K. McConathy, D. Duvick, and L.K. Mann. 1982. Effects of chronic air
pollution stress on photosynthesis, carbon allocation, and growth of white pine
trees. Forest Science 28: 60-70.
Miller, H.G. 1984. Dynamics of nutrient cycling in plantation ecosystems. In: Nutrition
of Plantation Forests, eds. G.D. Bowen and E.K.S. Nambiar. London: Academic,
pp. 53-78.
Miller, H.G. 1983. Studies of proton flux in forests and heaths in Scotland. Ini: Effects
of Accumulation of Air Pollutants in Forest Ecosystems, eds. B. Ulrich and J.
Pankrath. Boston, Massachusetts: D. Reidel, pp. 183-193.
39
Miller, P.R. 1973. Oxidant-induced community change in a mixed conifer forest. In:
Air Pollution Damage to Vegetation, ed. J. A. Naegele. Advanced Chemistry
Series 122. Washington, D.C.: American Chemistry Society, pp. 101-117.
Miller, P.R., J.R. Parameter, Jr., B.H. Flick, and C.W. Martinez. 1969. Ozone dosage
response of Ponderosa pine seedlings. Journal of the Air Pollution Control
Association 19: 435-438.
Monitor, A.V. and D.J. Raynal . 1983. Atmospheric deposition and ionic input in Adiron-
dack forests. Journal of the Air Pollution Control Association 33: 1026-1032.
Morrison, I.K. 1984. Acid rain. A review of literature on acid deposition effects in
forest ecosystems. Forestry Abstracts 45(8): 483-506.
Morrison, I.K. 1983. Composition of percolate from reconstructed profiles of two jack-
pine forest soils as influenced by acid input. In: Effects of Accumulation of
Air Pollutants in Forest Ecosystems, eds. B. Ulrich and J. Pankrath. Boston,
Massachusetts: D. Reidel, pp. 195-206.
Natr, L. 1975. Influence of mineral nutrients on photosynthesis and the use of assimi-
lates. In: Photosynthesis and Productivity in Different Environments, ed.
J. P. Cooper. IBP Synthesis Meeting on the Functioning of Photosynthetic
Systems in Different Environments. 1975 April; Aberystwyth, Wales. Cambridge:
Cambridge University Press, pp. 537-556.
Nieboer, E., D.H.S. Richardson, K.J. Puckett, and F.D. Tomassini. 1976. The phytotoxicity
of sulphur dioxide in relation to measurable responses in lichens. In: Effects
of Air Pollutants on Plants, ed. T.A. Mansfield. London: Cambridge University
Press, pp. 61-85.
Nihlgard, B. 1985. The ammonium hypothesis - an additional explanation to the forest
dieback in Europe. Ambio XIV(l): 2-8.
Noland, T.L. and T.T. Kozlowski. 1979. Effects of SO2 on stomatal aperature and sulfur
uptake of woody angiosperm seedlings. Canadian Journal of Forest Research
9(1): 57-62.
Norby, R.J., D.D. Richter, and R.J. Luxmoore. 1985. Physiologic processes in soybean
inhibited by gaseous pollutants but not by acid rain. New Phytologist 100:
79-85.
Oeschger, H., U. Siegenthaler, and M. Niemann. 1980. The carbon cycle and its pertur-
bation by man. In: The Interactios of Energy and Climate, eds. W. Bach, J.
Pankrath, and J. Williams. Boston, Massachusetts: D. Reidel, pp. 107-127.
Ormrod, D.P. 1982. Air pollutant interactions in mixtures. In: Effects of Gaseous Air
Pollution in Agriculture and Horticulture, eds. M.H. Unsworth and D.P. Ormrod,
London: Butterworth Scientific, pp. 307-331.
Overrein, L.N., H.M. Seip, and A. Tollan. 1980. Acid precipitation effects on forest
and fish: Final report of SNSF Project 1972-1980. SNSF Project FR 19/80.
Oslo, Norway, p. 104.
Paces, T. 1985. Sources of acidification in central Europe estimated from elemental
budgets in small basins. Nature 313: 31-36.
Patten, D.K. 1983. The effect of acid rain on the growth and nutrient content of two
species of hardwood tree seedlings, and on the pH, micrflora and nutrient
content of the soil. Ames, Iowa: Iowa State University. 178 pp. Ph.D. Thesis.
Pielou, E.C. 1982. Testing for vegetation gradients in the Athabasca Oil Sands area.
In: Acid Forming Emissions in Alberta and Their Ecological Effects, eds. H.S.
Sandhu, J.R. Clements, and B.L. Magill. 1982 March 9-12; Edmonton, Alberta,
pp. 435-448.
Postel, S. 1985. Protecting forests from air pollution and acid rain. In: State of the
World, 1985, ed. L. Brown. New York: W.W. Norton, pp. 97-123.
40
Postel, S. 1984a. Air Pollution, acid rain, and the future of forests. Part I. American
Forests 90(7) : 25-29.
Postel, S. 1984b. Air pollution, acid rain, and the future of forests. Part II. Ameri-
can Forests 90(8) : 12-16.
Postel, S. 1984c. Air pollution, acid rain, and the future of forests. Part III. Ameri-
can Forests 90(9) : 28-31 .
Postel, S. 1984d. Air pollution, acid rain, and the future of forests. Part IV. Ameri-
can Forests 90(10): 35-38.
Postel, S. 1984e. Air pollution, acid rain, and the future of forests. Part V. American
Forests 90(11): 46-47.
Postel, S. 1984f. Air pollution, acid rain, and the future of forests. Part VI. American
Forests 90(12): 38-39.
Raynal, D.J., J.R. Roman, and W.M. Eichenlaub. 1982. Response of tree seedlings to acid
precipitation - 1. Effect of substrate acidity on seed germination.
Environmental and Experimental Botany 22: 377-383.
Reid, D.M. 1985. Ethylene: A possible factor in the response of plants to air pollutants
and acid precipitation. I_n: Effects of Acid Deposition on Forests, Wetlands
and Agricultural Ecosystems, Presented at NATO Advanced Workshop, 1985
May 13-17; Toronto, Ontario. Heidelberg: Springer-Verlag (in press).
Rosenqvist, I. Th. 1978. Alternative sources for acidification of river water in Norway.
The Science of the Total Environment 10: 39-49.
Scheffer, T.C. and G.G. Hedgecock. 1955. Injury to northwestern forest trees by sulfur
dioxide from smelters. Forestry Service Technical Bulletin No. 1117. Washington,
D.C.: US Dept. of Agriculture. 49 pp.
Skarby, L. and G. Sellden. 1984. The effects of ozone on crops and forests. Ambio 13(2):
68-72.
Smith, W.H. 1981. Air Pollution and Forests. New York: Springer-Verlag. 379 pp.
Tabatabai, M.A. 1985. Effect of acid rain on soils. Critical Reviews in Environmental
Control 15(1): 65-110.
Tingey, D.T. and G.E. Taylor, Jr. 1982. Variation in plant response to ozone: A concep-
tual model of physiological events. In: Effects of Gaseous Air Pollution in
Agriculture and Horticulture, eds. M.H. Unsworth and D.P. Ormrod. London:
Butterworth Scientific, pp. 113-138.
Ulrich, B., R. Mayer, and P.K. Khanna. 1980. Chemical changes due to acid precipitation
in a loess-derived soil in central Europe. Soil Science 130(4): 193-199.
Van Breemen, N. 1985. Acidification and decline of central European forests: Nature
315: 16.
Van Ryne, D.M. and J.S. Jacobson. 1984. Effects of Acidity on Tree Pollen Germination
and Tube Growth. US Department of the Interior, Water Research Institute
Program, Project No. G859-04. 27 pp.
Van Ryne, D.M., J.S. Jacobson, and J. P. Lassoie. 1986. Effects of acidity on in vitro
pollen germination and tube elongation in four hardwood species. Canadian
Journal of Forest Research 16: 398-400.
Vitousek, P.M., J.R. Gosz, C.C. Grier, J.M. Melillo, W.A. Reiners, and R.L. Todd. 1979.
Nitrate losses from disturbed ecosystems. Science 204: 469-474.
Wellburn, A.R. 1982. Effects of SO2 and NO2 on metabolic function. in: Effects of
Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H. Unsworth and
D.P. Ormrod. London: Butterworth Scientific, pp. 169-187.
41
Wilhour, R.G. and G.E. Neely. 1977. Growth response of conifer seedlings to low ozone
concentrations. Iin: International Conference on Photochemical Oxidant
Pollution. Proceedings: Volume II, 1976 September 12-17; Raleigh, North
Carolina; US Enviromental Protection Agency Report EPA-600/3-77-001 6; pp.
635-645.
Woodman, J.N. 1987. Pollution-induced injury in North American forests: Facts and
suspicions. Tree Physiology 3: 1-15.
42
43
3. ACIDIC DEPOSITION EFFECTS ON AGRICULTURE
Agriculture and grazing are dominant in four ecoregions of Alberta: Short
Grass, Mixed Grass, Fescue Grass, and Aspen Parkland Regions, which cover about 25% of
Alberta (Strong and Leggat 1981). Thus, the effect of air pollutants on agriculture is
of both economic and ecological concern.
Agricultural production contributed 10.2% of Alberta's gross domestic product
in 1981, with grain crops such as wheat and barley accounting for over 75% of Alberta's
total farm cash receipts (Alberta Agriculture Statistics Branch, letter 1985). Almost
30% of Alberta's land area is used for farming, with 12% being cultivated at a given
time (Alberta Agriculture 1982). Approximately half (nearly 6 million hectares) of
Alberta's 12.5 million hectares of improved land is utilized for grain crops, including
barley (2.6 million hectares), wheat (2.7 million hectares), rye (.12 million hectares),
and mixed grains (.06 million hectares). Forages, including grasses and leguminous
forages, are farmed on approximately 1.8 million hectares. Oil-seed crops occupy .6
hectares. Other important crops are sugar beets (16,200 hectares), potatoes (8,000
hectares), field beans (3,600 hectares), and field peas (3,600 hectares).
The following discussion has been divided into three major sections. The first
section deals with the effects of wet acidic deposition on agriculture; the second
section discusses the effects of dry deposition; and the third section deals with the
effects of various mixtures of acid forming gases and others on agriculture.
3.1 EFFECTS OF ACIDIC PRECIPITATION ON CROPS
The effects of acidic precipitation on agricultural plants include reduction in
growth and yield, interference with reproduction, foliar injury, and alteration of foliar
processes such as fertilization, buffering, leaching, and nutrient accumulation. For
both ecological and economic concerns, plant growth, yield, and reproduction are the
most important plant responses. In the experiments reviewed by Torn et al. (1987),
dicotyledons appeared more likely to show inhibited growth than monocotyledons. None of
the experiments reviewed by these authors showed growth inhibition at steady state
simulated acidic rain with pH values above 4.0.
3.2 FOLIAR INJURY
Visible foliar injury (VFI), is the most frequently reported symptom of plant
response to simulated acidic precipitation. However, a direct correlation between
visible foliar injury and yield has yet to be established (Evans et al. 1981a, 1982c;
Johnston et al . 1981; Lee 1981; and Proctor 1983), and there is no known index of VFI
correlated with either growth or yield. Simulated acidic rain has, however, induced
visible injury on the foliage, fruits, and flowers of agricultural and horticultural
crop species. The literature on the topic indicates that no clearcut cause-effect
relationships exist; for example, there are cases where injured plants exhibited stimu-
lated growth, and others where growth was inhibited (Hindawi et al. 1980; Evans et al.
1981a) .
Foliar injury may reduce productivity through structural changes such as
inducing necrotic lesions or curling and wilting of the leaf, and/or through physiologi-
cal changes such as altering diffusive resistance or reducing intercellular spaces. The
44
indirect and direct effects of foliar injury attributed to acidic deposition are
summarized in Table 10 (Torn et al. 1987).
Most reports of VFI occur at or below pH 3.5. Sensitive species exposed to
rain events of pH 3.0 lasting two or more hours face a significant risk of foliar injury.
Current annual average pH of ambient rain in Western Canada is greater than 5 (U.S.
National Research Council 1983, Lau and Das 1985) and, therefore, poses little risk to
most field crops with respect to incurring VFI.
The preceding discussion, however, must be evaluated in the context of the
characteristics of ambient precipitation. The chemical composition of rainfall is never
constant. The chemistry of rain is known to vary within a given event (Pratt et al.
1983) and between events (Krupa et al. 1987). In addition, the frequency distributions
of ion concentrations in precipitation are skewed to the left (low concentrations) with
a long tail toward higher concentrations. These types of "non-normal" distributions
provide overestimations of the mean values for ions. Thus, experiments with "simulated
rain" with constant chemical composition derived from the mean of ambient precipitation
composition should be viewed as highly artificial and inappropriate in the context of
the real world situation.
The economic value of plants that sustain foliar or fruit damage can be greatly
reduced primarily because of appearance. However, actual yield appears not to be
affected by foliar injury for grains, forages, and processed fruits and vegetables.
Plants that exhibit injured leaf surfaces are likely to be more susceptible to
pathogen attack. Lesions formed on leaves by either wet or dry deposition can provide
infection sites for pathogens (Shriner and Cowling 1980). Evans and Curry (1979)
observed that in soybean, the vascular tissue induced depressions at the base of
trichomes and stomata in which acidic droplets accumulated and lesions were formed.
Penetration to the inner leaf structures by acidic solutions, possibly through micropores
and/or glandular hairs, is likely enhanced around both trichomes and stomata (Crafts
1961).
At the cellular level, foliar injury induced by acidic rain has been shown to
cause a reduction in mesophyll conductance, intercellular space, and the size of starch
granules (Ferenbaugh 1976; Neufeld et al . 1985a). Consequently, nutrient uptake and
carbohydrate storage may be affected, in turn affecting fruit set.
Wet acidic deposition affects leaf, surface tissues initially, whereas dry
deposition causes injury to internal cells (Evans et al. 1977; Evans and Curry 1979).
As a result of acidic, wet deposition, external lesions may eventually lead to internal
leaf cell injury (Evans et al. 1977; Hindawi et al. 1980). Necrotic lesions can also be
formed where entire cell strata are dead and bleached cells of chlorotic lesions exhibit
a reduced level of metabolism.
Acidic wet deposition has been found to form galls (elevated portions of leaves)
in bush bean, sunflower, wormwood, wax bean, and spinach (Adams and Hutchinson 1984).
Because of the shape of galls, once the initial injury is caused, further accumulations
of acidic droplets are prevented; thus galls actually reduce further injury to the plant
(Evans et al. 1977). The main effect of galls would likely be in reduced market value
of leafy commercial products because of their negative visual appeal.
Morphological changes in leaf tissue in response to simulated acidic wet
deposition, as noted, cause variations in the effects of a particular exposure to
Table 10. Potential effects of acidic precipitation on vegetation.
DIRECT EFFECTS
1. Damage to protective surface structures such as cuticle.
2. Interference with normal functioning of guard cells.
3. Poisoning of plant cells after diffusion of acidic substances
through stomata or cuticle.
4. Disturbances of normal metabolism or growth processes without
necrosis of plant cells.
5. Alteration of leaf- and root exudation processes.
6. Interference with reproductive processes.
7. Joint efforts with other environmental stress factors.
INDIRECT EFFECTS
1. Accelerated leaching of substances from foliar organs.
2. Increased susceptibility to drought and other environmental stress
factors.
3. Alteration of symbiotic associations.
4. Alteration of host/parasite interactions.
Adapted from Morrison (1984) and Tamm and Cowling (1976)
46
simulated acidic rain. Galls act as protective mechanisms and lesions formed around
stomatal openings also tend to protect the plant following a period of initial injury.
Stomata tend to close in the presence of acids (Plocher et al. 1985). Other depressions
formed as a result of acidic action on the leaf surface, for example, erosion of waxes
or cell collapse, can increase water capacity and retention times, thereby increasing
the probability of injury by increasing contact time with acids (Evans et al . 1977;
Jacobson and Van Lueken 1977; and Evans and Curry 1979). In other experiments cited by
the previous authors, acidic precipitation caused guard cells surrounding stomata to
increase in turgor, resulting in a reduction in diffusive resistance to gas exchange.
This effect may cause affected plants to be prone to water stress and wilting. Since
lower diffusive resistance can also result in higher photosynthetic uptake, the net
effect on plant response on productivity is not clear (Torn et al. 1987). Free hydrogen
ions may also affect internal cellular functions through reduction in ATP activity if
proton concentrations across membranes are changed. Exposure to acidic precipitation
has also been shown to cause chlorophyll levels in plants to decrease (both chlorophyll
a and b) (Ferenbaugh 1976; Hindawi et al. 1980; and Johnston et al . 1981).
The stage of plant development at the time of exposure to acidic deposition can
influence the nature of the response and this also varies by species. Foliar injury is
most pronounced on some species just prior to full leaf expansion (Evans 1984b). More
commonly, newly expanded and older pre-senescent leaves are the most susceptible to
acidic deposition and show the symptoms of VFI (Evans et al . 1977, 1981a; Evans and
Curry 1979; Keever and Jacobson 1983a; and Neufeld et al. 1985a). The effects of VFI,
however, are not clear. For example, in experiments with soybean exposed to simulated
acidic precipitation, Evans et al. (1977) found that leaves were least sensitive prior
to expansion but that newly expanded leaves were quite sensitive until 20 days old.
Significantly, these experiments showed that no injury or VFI was apparent on these same
plants at harvest, suggesting that the effect was not permanent or really detrimental to
either the plant or to its marketability.
Rapidly expanding leaves have incomplete wax coverage, leaving portions of the
cuticle exposed and, therefore, susceptible to acid injury (Neufeld et al. 1985b). The
cuticular waxes may act as a barrier preventing aqueous ions from penetrating the leaf
surface. The cuticle is also more hydrophilic than are the surface waxes; thus, expand-
ing leaves have a higher wettability and water retention than do unexpanded or fully
expanded leaves. In addition to these factors, shape of the emerging leaves can also
contribute to their higher wettability and consequential susceptibility to acid effects
(Neufeld et al . 1985b).
Wettability is a general term describing the amount of leaf surface area in
contact with and retention of a water droplet and it is positively correlated with the
degree of foliar injury but not necessarily with stress threshold (Keever and Jacobson
1983c). Susceptibility to foliar injury in relation to acidic deposition and the dosage
level necessary to cause visual effects and natural plant resistance to foliar injury
are not directly correlated or linked (Neufeld et al. 1985b). Threshold is defined as
the dose (in terms of concentration of acid) that is sufficiently high that any higher
dosage will result in a deleterious effect such as VFI. However, the concept of "a"
single threshold is controversial and is subject to debate. Variability in plant
47
response to pollutant stress exists at the population, genus, species and cultivar
levels. Further, plant response to a given pollutant is highly variable due to the
influence of a number of time oriented parameters such as: the presence or absence of
other pollutants, pathogens and pests; physical climatology; soil conditions and the
dynamics of plant growth itself. Thus, is is more appropriate to consider a range or
series of threshold values (Krupa and Kickert 1987).
Morphological leaf characteristics vary by species, cultivar, and stage of plant
development. This variation in leaf shape may partially account for the differences in
sensitivity to acid induced VFI found throughout the literature. Growth conditions
affect the leaf surface and also affect plant sensitivity to acidic precipitation
(Neufeld et al. 1985b). Agricultural plants in general have a lower resistance to foliar
injury when grown in controlled environments compared with field grown varieties of the
same species (Evans etal. 1981a, 1982b; Irving and Miller 1981; Cohen etal. 1982;
Troiano et al. 1982; and Keever and Jacobson 1983b). This last point is extremely
important when one considers that most of the experimental evidence indicating VFI as a
result of acidic deposition was compiled using simulated acidic rain in controlled
enviromental conditions.
An increase in acidity, frequency of occurrence, duration of exposure, and/or
number of simulated acidic rain events increases the extent and degree of foliar injury
(Evans and Curry 1979; Jacobson 1984). More frequent events allow less time for leaf
recovery and may prevent the leaf from drying, which in turn allows pathogens more time
to attack. The occurrence of injury is positively correlated with the length of time
that the leaf is wet, and negatively correlated with the number of days between exposures
(Irving 1983). By contrast, for gaseous pollutants or dry deposition, the occurrence of
injury increases as the concentration and duration of the exposure increase.
Extreme variation in the pH of acidic rain also appears to be an important
factor in production of foliar injury. Rain events with a varying pH giving a volume-
rated mean pH of 3.0 will tend to induce more foliar injury than wil.l the same number of
events with rain at constant pH 3.0 (Irving and Miller 1981; Johnston et al. 1981; and
Lefohn and Brocksen 1984). The reasons for these findings are not clear at this time.
3.3 SENSITIVITY OF PLANTS TO FOLIAR INJURY CAUSED BY WET ACIDIC DEPOSITION
Susceptibility to foliar injury from acidic deposition varies among species,
and among cultivars within a species. The relative sensitivity of 36 crop species was
analysed, based on data from 13 field and 14 controlled environment experiments where
plants were exposed to simulated wet acidic deposition (Table 11). The highest pH
resulting in VFI and the lowest pH applied without resulting in VFI for a variety of
plants are shown in Table 11. The range between the two pH values approximates a
threshold for foliar injury, for each cultivar tested, under the conditions when the
experiments were conducted. Crops grown in a controlled environment consistently
displayed a lower tolerance to acidic deposition than did field grown crops.
The dose at which 50% of the plants sustained significant VFI was pH 3.0. This
corresponds well with thresholds of pH 3.0-3.5 estimated by other investigators (Torn
et al. 1987). Simulated rain in the ambient mean pH range of 4.0 caused foliar injury
in 9% of the experiments, only one of which was conducted under field conditions. Below
48
Table 11. Visible foliar injury resulting from simulated wet acidic
deposition: pH threshold. (From Torn et al. 1987)
Species Highest pH Lowest pH Growth Reference
with with Conditions^
Foliar Injury No Injury
ROOT
Beet cv. Detroit Dark Red
4.0
5.6
C.E.
1
Beet
4.0
_
F
2
Carrot cv. Danvers
3.0
3.5
C.E.
1
Radish
4.2
5.6
C.E.
3
Radish
2.7
-
F
4
Radish
2.8
-
F
5
Radish cv. Cherry Belle
3.5
4.0
C.E.
1
LEAFY
Lettuce
3.1
4,0
C.E.
3
Lettuce, Bibb cv. Limestone
3.5
4.0
C.E.
1
Lettuce, head cv. Great Lakes
3.5
4.0
C.E.
1
Mustard green cv. Southern Giant
3.5
4.0
C.E.
1
Spinach cv. improved thick leaf
3.5
4.0
C.E.
1
Swiss chard cv. Lucullus
4.0
5.6
C.E.
1
Tobacco cv. Burley 21
3.5
4.0
C.E.
1
COLE
Broccoli cv. Italian green
3.5
4.0
C.E.
1
Cabbage
3.0
3.5
C.E.
1
Cauliflower cv. Early Snowball
3.5
4.0
C.E.
1
TUBER
Potato cv. White Rose
3.5
4.0
C.E.
1
LEGUME
Alfalfa cv. Honeoye
3.1
4.0
C.E.
3
Alfalfa cv. Honeoye
2.7
F
5
Alfalfa rv Vprnal
3 . 5
4.0
C.E.
1
Bean, bush
2.5
3.0
C.E.
6
Bean, bush
3.0
C.E.
7
Bean, bush
3.2
4.0
C.E.
8
Bean, kidney
3.2
C.E.
9
Bean, kidney
2.8
C.E.
10
Bean, kidney
2.7
F
5
Bean, pinto
3.0
4.0
C.E.
11
Bean, snap
2.6
F
12
Greenpea cv. Marvel
3.5
4.0
C.E.
1
Peanut cv. Tennessee Red
3.5
4.0
C.E.
1
Red clover cv. Kenland
3.5
4.0
C.E.
1
Soybean
2.9
C.E.
13
Soybean cv. Evans (G-O)
3.5
4.0
C.E.
1
Soybean cv. Hark(G-l )
3.5
4.0
C.E.
1
Soybean cv. Norman
3.5
4.0
C.E.
1
Soybean cv. OR-IO
4.0
5.6
C.E.
1
Soybean cv. Amsoy 71
3.3
4.1
F
continued
14
49
Table 11 . (Continued) .
Species Highest pH Lowest pH Growth Reference
with with Conditions^
Foliar Injury No Injury
LEGUMES (continued)
Soybean cv. Amsoy 71
2.7
3.1
F
15
Soybean cv. Davis
3.4
4.2
C.E.
16
Soybean cv. Davis
3.2
4.0
F
17
Soybean cv. Wells
-
3.0
C.E.
18
Soybean cv. Wells
-
2.8
F
19
Soybean cv. Wells
-
3.0
F
18
FRUIT
Apple blossom, Golden Delicious
3.0
4.0
F
20
Apple blossom, Mcintosh
3.5
-
F
21
Apple foliage
2.5
F
21
Apple foliage, Empire
2.5
F
21
Apple foliage. Golden Delicious
3.0
4.0
F
20
Apple foliage, Golden Delicious
-
2.5
F
21
Apple foliage, Mcintosh
-
2.5
F
21
Cucumber cv. 5116 Cresta
3.5
4.0
C.E.
1
Grape leaves
2.5
-
F
22
Green pepper cv. Calif. Wonder
4.0
5.6
C.E.
1
Strawberry cv. Quinalt
3.0
3.5
C.E.
1
Tomato cv. Patio
3.5
4.0
C.E.
1
FLOWER
Sunflower
3.2
-
C.E.
23
Zinnia flower petals
2.8
-
C.E.
24
Zinnia foliage
2.8
-
C.E.
24
GRAIN
Barley cv. Steptoe
3.0
C.E.
1
Corn cv. Golden Midget
3.0
3.5
C.E.
1
Corn cv. Pioneer 3992
3.0
F
25
Oats cv. Cayuse
3.0
C.E.
1
Wheat
2.7
C.E.
3
Wheat cv. Fieldwin
3.0
C.E.
1
BULB
Onion cv. Sweet Spanish
3.0
C.E.
1
FORAGE
Bluegrass cv. Newport
4.0
5.6
C.E.
1
Fescue cv. Alta Tall
3.5
4.0
C.E.
1
Orchardgrass cv. Potomac
3.5
4.0
C.E.
1
Ryegrass cv. Linn
3.5
4.0
C.E.
1
Ryegrass, perennial
3.0
C.E.
Timothy cv. Climax
3.5
4.0
C.E.
1
^ C.E. = Controlled Environment
F = Field Grown
= Information not available
continued...
50
Table 11 . (Concluded) .
References: 1. Lee et al. (1981)
2. Evans et al. (1982a)
3. Evans et al . (1982c)
4. Troiano et al. (1982)
5. Evans et al. (1982b)
6. Ferenbaugh (1976)
7. Hindawi et al. (1980)
8. Johnston et al . (1981 )
9. Shriner (1974)
10. Paparozzi (1981)
11. Evans et al. (1980)
12. Troiano et al . (1984)
13. Evans and Curry (1979)
14. Evans et al. (1983)
15. Evans et al. (1981c)
16. Norby and Luxmoore (1983)
17. Brewer and Heagle (1983)
18. Irving and Miller (1981)
19. Troiano et al . (1983)
20. Proctor (1983)
21 . Forsline et al . (1983b)
22. Forsline et al . (1983a)
23. Jacobson and Van Leuken (1977)
24. Keever and Jacobson (1983a)
25. Plocher et al. (1985)
51
pH 2.5, 70% of the cultivars showed foliar injury. As previously stated, it is important
to note that in many of these experiments the use of "simulated rain" with constant
chemical composition derived from the mean values of precipitation pH and chemical com-
position, should be considered as highly artificial and unrelated to ambient conditions.
The groups of crops most susceptible to visible injury were, from most to least
susceptible, root, leafy, cole, legume, fruit, grain, and leafy and seed forage crops
respectively. The potential for economic loss was greatest for leafy, cole, and fruit
crops. Leafy crops showed slightly less vulnerability to acidity induced foliar injury
than did root crops. The threat to the economics of yield is, however, greater with the
leafy crops which may lose commercial value if blemished. The threshold for injury to
cole foliage is pH 3.0 to 3.5 which is higher than for leafy crops. Sensitivity of
legume species varied, with plants such as soybean tending to be most susceptible to
simulated acidic rain, likely due to their higher wettability. Foliar injury was
observed for almost all fruit species studied. In addition, perennial fruit trees have
shown latent foliar injury after cessation of acidic rain treatments. However, growth
of annual fruit crops is, in general, stimulated by simulated acidic rain. Monocots,
such as wheat, barley, and timothy were found to be resistant to foliar injury when
exposed to simulated acidic rain above pH 2.5 (Torn et al. 1987).
3.3.1 Direct Foliar Effects of Wet Acidic Deposition
The plant leaf is composed of permeable tissue with continuous exchange of
gases, water, and dissolved substances. Foliage may react chemically with acidic
solutions upon contact without sustaining any change in its physical structure. Foliar
fertilization, buffering, and leaching are all processes that have been investigated in
this regard with respect to the effects of acidic deposition on plants. The acidic
solution may represent wet or dry deposition that has hydrolyzed on the leaf's surface.
These types of processes can result in direct but subtle effects on a plant because they
may be active on internal processes.
3.3.1.1 Foliar Fertilization. Acidic deposition can act as a source of the nutrients
nitrogen and sulphur that become available to plants if absorbed by the leaf (Irving and
Miller 1981; Troiano et al. 1983; and Evans 1984). This process, termed foliar fertili-
zation, can be both beneficial and detrimental to plants. Direct application of nutrients
in this manner is a fast way of supplying nutrients to leaves although transfer of the
fertilizer away from the sites of entry is a slow process (Garcia and Hanway 1976).
Nutrients applied to the foliage pose the risk of inducing foliar injury (Neumann et al.
1981). Although there are insufficient data for a definitive conclusion, it is widely
assumed that little of the nutrients contained in ambient precipitation penetrate the
foliage to any degree (Evans et al. 1981, 1983; Evans 1984b). Even commercially available
foliar fertilizers, which use added surfactants to aid foliar penetration, have had
limited success in stimulating plant growth (Torn et al . 1987). Reductions in yield
often correlated with foliar injury from fertilizer salts are often cited to be the
result of nutrient application at high concentrations. Neumann et al. (1981) concluded
that all osmotically active fertilizer compounds can induce plasmolytic damage when at
sufficiently high concentrations to penetrate the leaf. In fact, fertilizer doses small
52
enough to prevent foliar injury may not allow penetration of enough fertilizer to
stimulate plant growth (Neumann et al. 1981).
In their research on acidic rain, Evans et al. (1983, 1984) and Irving and
Miller (1981) have considered the effects of the nutrients being added in precipitation.
Evans et al. (1983) applied simulated acidic rain at pH 2.7 to soybeans. The plants
were thus exposed to 10 times the ambient atmospheric levels of N and S. The net effect
of this treatment was a 23% reduction in seed yield. It would appear that any fertili-
zation effect of the acidic deposition was not sufficient to offset the detrimental
effects of acidity.
Because the concentration and total deposition of nitrogen and sulphur in acidic
precipitation are far lower than those found in commercial foliar fertilizers, it appears
unlikely that significant benefit to crops will be realized in the form of foliar
fertilizer (Evans et al. 1981a). This is particularly true when one considers that
even the commercial foliar fertilizers have had limited success in improving plant growth
except at particular times and for short periods.
3.3.1.2 Foliar Buffering. Some plants appear to develop little or no foliar injury
from acidic precipitation. It is possible that the plant tissue may effectively buffer
the acid before any significant physical or physiological injury can occur, and this
ability may differ among species (Craker and Bernstein 1984).
Craker and Bernstein (1984) investigated the buffering ability of red kidney
bean, wheat, red clover, soybean, corn, and timothy by soaking leaf tissue in simulated
acidic rain solutions (pH 2.0, 3.0, or 4.0). In each case the pH of the solution rose
within four hours. Subsequent visual analysis of leaf tissue injury suggested that the
leaves with greater buffering capacity were more susceptible to foliar injury. Adams
and Hutchinson (1984) found that the ability of the leaf to buffer simulated acidic
precipitation was directly correlated with the extent of injury sustained. Foliar
leaching of potassium associated with exposure to simulated acidic rain may be a secon-
dary effect of foliar injury and may account for part of the buffering capability (Keever
and Jacobson 1983a). These results support the hypothesis that the buffering ability of
leaves is due to the release of cellular materials from dead or disrupted cells.
Senescent leaves have a much greater buffering capacity than young healthy leaves.
Bicarbonate stored in cell walls for photo'synthetic activity may act to neutralize
acidic deposition (Oertli et al . 1977). This should also mean that leaf litter layers
have a high capacity to counteract the effects of acidic deposition.
It is also possible that leachates or superficial aggregates of particulate
matter contribute to the buffering of excess hydrogen. For example, foliar alkaline
deposits formed from foliar leachates and atmospheric CO2 can neutralize acidic solu-
tions (Adams and Hutchinson 1984). In another investigation, surface contaminants and
microflora were removed from leaves prior to treatment with acidic rain with no effect
on the buffering ability of the leaves (Craker and Bernstein 1984).
3.3.1.3 Foliar Leaching. Foliar leaching has been studied by numerous investigators by
examining the chemical constituents of leachate following leaf exposure to simulated
acidic solutions (Evans et al. 1977; Hindawi et al. 1980; Keever and Jacobson 1983a, b,c;
53
and Adams and Hutchinson 1984). The results of these studies showed that leaching rates
of Ca, K, and Mg for seven out of nine plant species tested increased after exposure to
simulated acidic rain. Johnston et al. (1981), however, found both increased and
decreased leaching of foliar K in soybean subjected to simulated acidic rain. In other
studies on soybean conducted by Hindawi et al. (1980), no effect on the leaching of K
was found but levels of both Ca and Mg increased, as did P and NOa. In the previously
cited work by Johnston et al. (1981), these ions were not measured. On the basis of
these studies it would appear that simulated acidic precipitation does cause the leaching
of cations from leaves. Some of the possible mechanisms are discussed below.
Foliar potassium was found to increase in concentration in the leachate of
zinnias and soybean following exposure to acidic solutions (Keever and Jacobson 1983a, c).
In the case of zinnia, no effect on leaching was found at pH 4.0 or at the control
pH 5.6; however, at pH 2.8, K was found to increase in the leachate as measured by Rb86.
This loss of K from plant leaves was further accelerated under nutrient rich conditions.
An increase in leaching of foliar K associated with foliar injury was also found in
soybean and bean at a threshold pH of 4.0 by Keever and Jacobson (1983c) and Evans
et al. (1981), respectively. The threshold for leaching was of the same order of
magnitude as that observed for foliar injury in soybean (Evans et al. 1983c; Keever and
Jacobson 1983c). The foliar loss of potassium may have been due to the death and
subsequent degradation of cells resulting from exposure to a low pH solution (Keever and
Jacobson 1983c) .
Foliar buffering and increases in leaching due to acidity are undoubtedly
related processes. Buffering on the leaf surface is aided by alkaline deposits formed
by atmospheric deposition or by exuded foliar salts. Leaching occurs as exchangeable
cations in the cuticle and cell walls and these cations are exchanged for in acidic
solutions (Adams and Hutchinson 1984). The cuticle forms a barrier for ion movement in
and out of the tissue, and the cuticle waxes play a role in inhibiting leaching of
foliar nutrients (Neufeld et al. 1985b). Cuticular micropores are the principal route
for cation exchange and loss, as well as for entry of chemicals into the intercellular
structures of leaves (Adams and Hutchinson 1984; Evans 1984). Greater wettability is
correlated with both higher leaching and higher buffering capacity.
3.3.1.4 Foliar Nutrient Content. Increased foliar leaching may alter the nutrient
content of leaf tissue. A significant reduction of foliar N, P, Mg, and Ca was observed
in soybean leaves exposed to acid mist (Hindawi et al. 1980). Potassium content was not
affected while that of S increased in these experiments. Experiments with simulated
acidic rain at lower pH and soybean showed that foliar N and S increased, while Mn
decreased (Brewer and Heagle 1983). The inconsistency in these two studies is typical
of research in this subject area. However, the results in general, may be summarized as
follows: leaching rates of micronutrients increase, there is no net effect on foliar S,
and results vary for N (Torn et al . 1987). For example, the nutrient content of kidney
bean and soybean foliage when exposed to simulated acidic rain applications of pH 6.0
and 3.2 demonstrated a pH-independent response (Shriner and Johnston 1981). Other
studies suggest that detected effects may only be temporary.
54
It is suspected that energy diverted from growth may be the penalty plants pay
to replace leached metabolites when exposed to simulated acidic precipitation (Amthor
1984). Nutrient reductions may also affect nutritional quality of the plant and hence,
its economic value (Evans et al. 1981a; Evans 1984). In highly managed environments
such as propagation beds or container nurseries where root systems are either limited or
restricted, foliar leaching may lead to nutrient deficiency symptoms. However, this is
not likely to occur with field crops.
3.4 EFFECTS OF WET ACIDIC DEPOSITION ON PLANT GROWTH
Plant growth may be stimulated, inhibited, or not affected by exposure to wet
acidic deposition. The mechanisms by which wet acidic deposition alters plant produc-
tivity have as yet not been established. Using dose-response data, a qualitative ranking
of plant growth sensitivity to simulated acidic deposition has been prepared from the
current literature and is shown in Table 12 (Torn et al. 1987).
Although artificial wet acidic deposition has been shown to affect plant growth
under controlled conditions, no experiments or documentation of field effects on growth
were found that showed that ambient rates of acidic deposition negatively affect growth
(Torn et al . 1987). The growth of many species is stimulated or not affected by
simulated acidic rain in the ambient range. When the acidity dose exceeds a plant's
threshold, yield of the whole plant or some portions of it is decreased. An intermediate
effect between the threshold and control pH has been observed whereby the plant growth
may be increased. Lee (1981) showed this intermediate effect in seed germination,
seedling growth, and crop yield between pH 3.5 and 4.0. Although dose-response functions
for crop yield and quality are considered an aid in predicting impacts of ambient and
anticipated levels of acidity in rainfall (Troiano et al. 1982; Evans et al. 1983,
1984), the lack of linearity of such responses and the lack of understanding of their
mechanisms, limits the value of the current information base.
The results of studies on the effects of simulated acidic precipitation on yield
for a variety of agricultural crops are summarized in Table 12. The results clearly
show no effects under field type situations except in the case of beets (Irving 1983).
Root crops such as beets are the most sensitive agronomic group with low threshold and
resistance for both foliar injury and yield reduction (Torn et al. 1987). All other
field grown crops showed a growth peak at an intermediate pH treatment. Torn et al.
(1987) concluded in their review that there was no statistical significance in the
ranking of the sensitivity of agricultural crop species to increasing acidity, but that
there was evidence for a decline in yield of most species at exposures to acid mist
below pH 3.5, at a rate of 1 to 9% per pH unit decrease in the mist. However, exceptions
were orchard grass, timothy, and possibly bluegrass, and forage crop species where yield
increased between 2 and 24%. When these deviations were removed from the standardized
data set examined by Torn et al . , the percentage yield change for the remaining studies
was -3 (± 4). Since the standard deviation is higher than the mean, the validity or
reliability of the results may be questionable. The calculation was based on data from
studies on legumes, forage, and grain species (Torn et al. 1987 citing Lee and Neely
1980; Lee et al. 1981; Evans et al. 1982c; and Harcourt and Farrar 1980).
55
Table 12. Effect of simulated acidic rain on marketable yield of roots
and shoots.
Species Marketable Yield Response
to Increased Acidity
Growth
Conditions
References
ROOTS
Radish cv. Cherry Belle
no effect
r
1
1
Radish cv. Scarlet Knight
no effect
F
1
Radish
no effect
CE
2
Radish
decrease
CE
3
Radish
no effect
F
3
Beet
decrease
F
4
Beet
decrease
F
3
beet cv.uexroii uarK Kea
decrease
CE
5
Carrot cv.Danver's
decrease
CE
5
LEAFY
Lettuce
decrease
CE
3
Mustard green
decrease
CE
5
Lettuce, Bibb
decrease
CE
5
Lettuce, head
decrease
CE
5
COLE
Broccoli
decrease
CE
5
Caul i f 1 ower
no effect
CE
5
Cabbage
no effect
CE
5
TUBERS
Potato cv. Russet
no effect
F
Potato cv. Kennebec
no effect
F
1
Potato cv. White Rose
decrease
CE-
5
LEGUME
Alfalfa cv.Honeoye
decrease ^
CE
3
Alfalfa cv.Honeoye
no effect
1
Alfalfa cv. Vernal
no effect
F
6
Alfalfa cv. Vernal
increase
CE
5
FORAGE
Ryegrass
decrease 2
CE
7
Fescue cv.Alta
decrease 2
F
1
CE = Controlled Environment
F = Field
^Decrease 1 harvest; no effect 2 harvests.
^Decrease after only 3 or 4 harvests
continued .
56
Table 12. (Concluded) .
References: 1. Plocher et al . (1985)
2. Harcourt and Farrar (1980)
3. Evans et al . (1982a)
4. Troiano et al. (1982)
5. Lee et al. (1981)
6. Evans et al. (1982c)
7. Amthor and Bormann (1983)
57
3.5 EFFECTS OF WET ACIDIC DEPOSITION ON PLANT REPRODUCTION
Very little information is available in the current literature regarding the
effects of acidic deposition on seed germination, seedling emergence, pollen viability,
or fruiting (Torn et al. 1987). Seedling emergence of some woody species has been
reported to be inhibited, stimulated, or unaffected by acidic precipitation (Cox 1983;
Evans 1984b). Dilute acids, on the other hand, can have a scarifying effect on seed
coats, thus aiding germination (Morrison 1984). Among plant species, acidity has been
shown to inhibit in vitro germination of pollen of apple, grape, tomato, and camellia
plants (Kratky et al. 1974; Masaru et al. 1980; and Forsline et al. 1983a). While there
are no surveys of agricultural crops, estimates based upon forest research suggested a
threshold for inhibition of pollen germination in trees to be pH 3.6 (Cox 1982).
Comparison of foliar injury relationships suggests that pollen germination of agricul-
tural species may be more sensitive to acidic precipitation in comparison with trees
(Torn et al. 1987).
Acidic deposition may interfere with successful reproduction at different
seasons and/or at different stages of development for perennial species such as fruit
trees. Air pollutants may affect the fruiting process at the time of flower initiation
during the first year because the inflorescence can be very vulnerable to external
influences. Flowering in such plants coincides with periods of rainfall with high
acidity in many regions (Forsline et al. 1983b). Alterations in the bloom can influence
pollen germination and seed or fruit set, although mechanisms and responses have not
been documented at this time. During the second year, air pollutants may influence
fertilization, fruit set, fruit development, and maturation (Torn et al. 1987). Insect-
plant interactions and pollination may also be affected due to deformed flower structure.
In general, the effects of acidic precipitation on plant reproduction are still not
fully understood. The effects of acidic deposition on the sexual reproduction of corn,
wheat, snapbean, soybean, and other crops are currently under study at North Carolina
State University by DuBay and Stucky. However, no results are available at this time.
Fruits and flowers are highly susceptible to injury, and generally sustain
injury at lower levels of acidity compared with foliage (Jacobson and Van Leuken 1977;
Forsline et al. 1983b; Proctor 1983; and Keever and Jacobson 1983a). Blemished fruit,
if sold directly to consumers, generally has a lower market value. However, little or
no change in economic value can be attributed to injury if fruit is sold for canning or
extracting juice (Lee 1981). Presently there are no data relating injury on reproductive
structures to alteration in reproductive potential (Torn et al. 1987).
3.6 EFFECTS OF DRY DEPOSITION ON AGRICULTURAL CROPS
The following components of dry deposition will be discussed in this section:
SO2, NOx, Oa, and H2S.
Sulphur dioxide (SO2) is one of the two major acid forming air pollutants in
industrial emissions. It is very phytotoxic both in gaseous form and in its hydrated
form when dry deposition dissolves on wet plant parts. Sulphur dioxide is extremely
soluble in water under high pH. Susceptible plants may be injured by 0.05 to 0.5 ppm of
sulphur dioxide after exposures as short as eight hours (Mudd and Kozlowski 1975).
58
Nitrogen oxides, particulary NO2 and NO, usually are only present at phyto-
toxic levels in severely polluted environments. With time, the NO2 level tends to
decrease because of photochemical transformation processes that lead to ozone (O3)
production. Because of these reactions, phytotoxic levels of NO2 are not of great
concern to agricultural interests. Continuous exposure to 0.25 to 0.5 ppm of NO2 can
cause VFI in sensitive plants (Taylor and Maclean 1970; National Academy of Sciences,
U.S. 1977b).
On the other hand, ozone is very phytotoxic and research into its effects on
agricultural crops and plants in general has been carried out for the past thirty years.
Exposure of very sensitive plants to ozone at concentrations as low as 0.10 ppm for one
hour, or long term average concentrations of 0.03 ppm with periodic or intermittent
episodes can be detrimental to foliage, growth, and yield. Ozone exposure of plants of
intermediate sensitivity will induce injury at concentrations of 0.30 ppm for one hour,
or 0.10 ppm for several hours (Guderian 1985). The threshold concentration for sensitive
cultivars with respect to chronic ozone exposure has been set by the National Academy of
Sciences, U.S. (1977a) as 0.05 to 0.1 ppm.
Phytotoxic levels of hydrogen sulphide have been found to be well outside the
ambient range and present levels do not likely pose a threat to agricultural plant
species (Heck et al. 1970). Concentrations as high as 0.3 ppm generally have no adverse
effects on plants and can even stimulate growth (Torn et al . 1987). As opposed to other
acid precursor gases, hydrogen sulphide can cause more injury in drier soils than under
wet conditions (Thompson and Kats 1978). Because of its general lack of effect at
current ambient levels, H2S will not be discussed further in this synthesis.
3.6.1 Physiological Effects of Dry Deposition
3.6.1.1 Sulphur Dioxide Effects on Stomata. Sulphur dioxide directly affects the
stomata, which may be induced to open or close depending on plant species, pollutant
concentration, duration of exposure, and prevailing environmental conditions.
Sulphur dioxide induced stomatal opening has been observed in several plant
species under fumigation conditions: field bean, corn, pine, bush bean, navy bean,
white bean, pea, grapevine, radish, sunflower, tobacco, cucumber, soybean, and two
varieties of saltbush (Torn et al. 1987). Stomatal opening occurred within a few minutes
of sulphur dioxide fumigation and resulted in a 10% to 20% increase in stomatal conduc-
tance in several four-carbon (C4) species and increases as high as 200% in three-carbon
species (Black 1982).
The opposite effect, stomatal closure as a result of sulphur dioxide fumigation,
has been detected in a wide variety of plants. The stomata of the following tree species
were induced to close under fumigation: pine, poplar, birch, and apple. A number of
important agricultural species were also affected in this manner. Stomatal closure has
the effect of inhibiting transpiration. The maximum transpiration inhibition rate
observed in these various tree and agricultural species ranged from 35% to 75% and
occurred within ten minutes to four hours following exposure, depending on the species
examined (Black 1982) .
59
The majority of the aforementioned fumigation experiments were conducted using
concentrations of sulphur dioxide higher than those found in polluted environments. It
is not known whether these species would show similar responses at more realistic con-
centration levels. Ziegler (1975), from work in polluted environments, has consistently
observed increases in stomatal conductance and transpi rational losses as a result of
sulphur dioxide exposure. The initial increase was 15 to 20% followed by a decrease in
transpiration of up to 50% in the species she studied. Ziegler further stated that low
concentrations of sulphur dioxide can cause a permanent increase in transpiration.
Whether the increased stomatal aperture during these exposures is caused by increased
turgidity of the guard cells, a reduction in turgidity within the epidermal cells
adjacent to the guard cells, or other mechanisms is as yet undetermined (Black 1982).
Once the sulphur dioxide enters the leaf through the stomata, it reaches the
mesophyll cells where it is hydrolyzed in the surface fluid to become sulphite. The
buffering capacity of cytoplasm decreases with time under acid conditions and especially
with an increased sulphur dioxide concentration. Sulphite is toxic and therefore, is
mostly oxidized to sulphate and stored. These sulphates are later converted to organic
sulphur compounds or exuded by the roots. Sulphate accumulations, primarily at the
edges and tips of leaves, increase with increased photosynthesis and are therefore at
the maximum in young leaves. If the plant's capability to oxidize sulphites is exceeded,
sulphites can build up to toxic levels and result in injury to the plant (Ziegler 1975).
3.6.1.2 Sulphur Dioxide Effects on Photosynthesis. Most studies indicate a decrease in
photosynthesis with increased sulphur dioxide exposure (Mudd and Kozlowski 1975; Black
1982). Depression of photosynthesis occurs quickly and is readily reversible if the
sulphur dioxide concentration drops. Responses are less reversible at higher con-
centrations. The lowering of the photosynthetic rate appears to be associated with
breakdown of biochemical systems, tissues, and the appearance of visible foliar injury.
Sulphur dioxide induced changes in photosynthesis are also influenced by
irradiance and temperature. It is hypothesized that these factors may influence the
rates of detoxification or biochemical processes (Black 1982).
3.6.1.3 Sulphur Dioxide Effects on Respiration. Some investigators have shown an
inhibitory effect on dark respiration as a result of exposure to high concentrations of
sulphur dioxide (Gilbert 1968; Taniyama 1972). However, other investigators found
stimulatory effects (Keller and Muller 1958; De Koning and Jegier 1968; Taniyama et al.
1972; Baddeley and Ferry 1973; and Black 1982). Similar conflicting results were found
in studies on the effects of sulphur dioxide on photorespi ration (Ziegler 1975; Koziol
and Jordan 1978; and Black 1982). With these conflicting results it is not possible to
derive a general conclusion on the effects of sulphur dioxide fumigation on respiration.
3.6.1.4 Nitrogen Oxide Effects on Stomata and Transpiration. There are few available
data on the direct effects of nitrogen oxides on plant stomata. Indirect effects were
reported by Hill and Bennett (1970) who found that NOx inhibition of photosynthesis
resulted in a carbon dioxide buildup in intercellular spaces, causing the stomata to
close.
60
After entering the plant through the stomata, nitrogen oxides diffuse through
the intercellular spaces to the mesophyll and parenchyma where they react with the
hydrated cell surfaces to form a mixture of nitrous and nitric acids. When these acids
exceed threshold values they may cause injury to tissue (Mudd 1973; Zeevaart 1976; and
McLaughlin et al. 1979).
3.6.1.5 Nitrogen Oxide Effects on Photosynthesis. Hill and Bennett (1970) found in
studies on alfalfa and oats, that a concentration of 0.6 ppm of NO or NO2 reduced
carbon dioxide assimilation (a measure of photosyntheti c activity). Conversely, the
gases in combination showed an additive effect causing photosynthesis to be lowered
further. Of the two gases, NO2 appeard to affect photosynthesis less and allowed
faster recovery than NO. Increases in photosynthesis have also been observed at low
level nitrogen oxide fumigations, likely as a result of a fertilizer effect (Bull and
Mansfield 1974).
3.6.1.6 Nitrogen Oxide Effects on Respiration. There are no data available concerning
the direct effects of nitrogen oxides on plant respiration.
3.6.1.7 Ozone Effects on Stomata. Transpiration, and Photosynthesis.
Ozone is believed to increase the permeability of cell membranes and cause
leakage of ions. Intercellularly, ozone can attack the plasmalemma of inner cell walls.
This causes the permeability of the lining to be disrupted, allowing the leakage of cell
contents into the intercellular spaces (Wedding and Erickson 1955; Perchorowicz and Ting
1974). Without entry to the cellular spaces, most researchers feel that ozone does not
affect plants. However, experimental data on this point are contradictory.
Most researchers agree that ozone can induce stomatal closure in plants. This
has the effect of inhibiting transpiration. Stomatal closure also in turn can contribute
to the resistance of the plant to ozone injury (Engle and Gabelman 1966; U.S. EPA 1978).
It is generally accepted that ozone inhibits photosynthesis and that this
inhibition can occur without foliar injury (Tingey 1977; U.S. EPA 1978). In addition,
ozone alters the way in which the products of photosynthesis are distributed within
plants ( Jacobson 1982) .
3.6.2 Foliar Effects of Dry Deposition
The most readily observed symptoms of gaseous pollutant exposure on plants are
visible foliar injury. Foliar effects can be divided into two categories: acute and
chronic .
Acute injury to plant tissue occurs within hours or days after exposure to
short-term (less than 24 hours), high concentrations of pollutants. Chronic injury on
plants usually develops after long-term exposure to variable, but lower, concentrations
of the pollutants, with periodic, intermittent episodes.
Foliar injury caused by SO2, NOx, and H2S is usually found in areas near
emission sources. Conversely, foliar injury due to ozone is often found on a regional
scale, downwind from industrial and urban sources.
61
3.6.2.1 Foliar Effects of Sulphur Dioxide. Acute injury caused by sulphur dioxide is
usually found as foliar necrosis in which metabolic processes cease and plant cells are
killed. Chlorosis may also be observed. Chronic injury includes chlorosis in which the
cells are not killed, but chlorophyll is converted to phaeophytin and leaves become
bleached. The leaves remain turgid but function less efficiently (Linzon 1978).
Acute injury from sulphur dioxide exposures is caused by a rapid accumulation
of bisulphite and sulphite (Linzon 1978). When the oxidation product, sulphate, accumu-
lates beyond a threshold value that the plants can tolerate, chronic injury also occurs.
Linzon (1978) estimated that sulphate is about 30 times less toxic than sulphite.
Chronic foliar injury is typified by yellowing or bronzing which may occur due
to the presence of pigments previously masked by chlorophyll that has been destroyed.
Chlorosis in chronic injury is generally interveinal on broad leafed plants (Torn et al.
1987).
These visual symptoms are characteristic of sulphur dioxide induced foliar
injury but they can only be used as a guide in identifying the cause of injury because
other factors influence plant injury as well, such as climate, insects and other pests,
soil nutrition, and genetic and physiological factors. Table 13 indicates the threshold
concentrations of sulphur dioxide that induce visible foliar injury for various species.
The threshold SO2 concentrations for VFI range from of 0.18 ppm for eight hours to
2.0 ppm for one hour.
The sensitivity of agricultural crops to sulphur dioxide are summarized in
Table 14. Sensitivity was based on VFI with sulphur dioxide exposures under conditions
favourable for gas absorption by plants (Barrett and Benedict 1970).
3.6.2.2 Foliar Effects of Nitrogen Oxide. Nitrogen dioxide is the only oxide of
nitrogen that has been found to injure vegetation at concentrations found in ambient air
but under very select conditions. Even when controlled fumigations of NO were conducted,
visible symptoms were not seen at concentrations as high as 25 ppm (Legge et al. 1980).
The middle-aged to oldest leaves were most susceptible to injury, although this varied
by species.
The most commonly observed symptoms of acute nitrogen dioxide injury on broad
leafed plants are interveinal water-soaked lesions on the adaxial leaf surface which
appear one to two hours following exposure. These lesions rapidly collapse and bifacial
necrotic areas develop. These areas are bleached to a white, light tan, or bronze
colour when dry. Lesions gradually extend through the leaf to produce small irregular
necrotic patches (Torn et al. 1987). This injury is similar to that seen as a result of
sulphur dioxide exposure (Taylor and MacLean 1970; Taylor et al. 1975). In sensitive
species, lesions occur at the margins and at the apex of leaves (Taylor and MacLean
1970). Acute injury generally occurs at nitrogen dioxide concentrations of between 1.6
to 2.6 ppm or greater for exposures of up to 48 hours (Legge et al. 1980).
Symptoms of chronic nitrogen dioxide injury include chlorosis and premature
defoliation and fruit drop. An enhancement of the green colour may be observed prior to
the onset of these symptoms (Taylor and MacLean 1970; Legge et al. 1980).
Van Haut and Stratmann (1967) fumigated 60 species of plants with a onerone
mixture of NO and NO2. On the basis of their results, a classification of Alberta
agricultural plants as to their relative sensitivity is provided in Table 15.
62
Table 13. Threshold sulphur dioxide concentrations (ppm) causing
foliar injury to various agricultural species.
Exposure Time
A. Field observations
Dreisinger and McGovern (1970) Ih 2h 3h 4h 8h
(Ni/Cu smelters -
Sudbury, Canada)
Sensitive crops
0,
.70
0.
.40
0.
34
0.
.26
0.
18
Intermediate
0.
.95
0,
.55
0.
43
0.
.35
0.
24
Resistant
1 ,
.88
1 .
.1
0.
86
0,
.70
0.
49
Jones et al. (1979) 1 h 3 h
(Power plants -
Tennessee, US)
Sensitive 0.50 to 1.0 0.30 to 0.60
Intermediate 1.0 to 2.0 0.60 to 0.80
Resistant 2,0 + 0.80 +
B. Controlled environment
fumigations
Van Haut and Stratmann Ih 2h 3h 4h 8h
(1967)
Sensitive (rye) 2.3 1.9 1.1 - 0.75
Katz and Ledingham (1939)
Sensitive
(alfalfa, barley) 1.5 1.0 0.89 - 0.55
continued.
63
Table 13. (Concluded) .
Exposure Time
B. Controlled environment
fumigations (continued)
Thomas (1935) Ih 2h 3h 4h 8h
Sensitive
(alfalfa) 1.2 0.71 0.55 0.48 0,36
Fujiwara (1975)
Sensitive - 0.60 0.45 - 0.25
Zahn (1961)
Sensitive
0.70
0.62
0.60
0.58
0.50
Intermediate
1.2
1 .1
1 .0
1.0
0.9
Resistant
1 .8
1 .7
1.6
1.6
1 .4
Adapted from the original table in International Electric Research
Exchange (1981).
64
Table 14. Agricultural species sensitive to sulphur dioxide. ^
Alfalfa
(Medicaqo sati va)
Barley
(Hordeum vulgare)
Bean, field
( Phaseolus vulgaris)
Beet, table
(Beta vulgaris)
Broccoli
(Brassica oleracea cv. botrytis)
Brussel sprouts
(Brassica oleracea cv. gemmif era)
Carrot
(Daucus carota var. sati va)
Clover
(Meli lotus & Trifolium sp.)
Cotton
(Gossypium sp. )
Lettuce
( Lactuca sati va)
Oats
(Ayena sati va)
Radish
(Raphanus sati vus)
Rye
(Secale cereale)
Saf flower
(Carthamus tinctorius)
Soybean
(Glycine max)
Spinach
(Spinacea oleracea)
Squash
(Cucurbita maxima)
Sweet Potato
( Ipomoea batatas)
Swiss Chard
( Beta vulgaris cv. cicla)
Turnip
( Brassi ca rapa)
^Sensitivity is based on foliar injury
Source: Barrett and Benedict (1970)
65
Table 15. Suggested susceptibility of various agricultural species
which occur in Alberta to a combination of nitrogen dioxide
and nitric oxide.
Plant Species
Sensitivity Category^
Alfalfa
(Medicaqo sativa)
Barley
(Hordeum distichon)
Crimson or Italian clover
(Trifolium incarnatum)
Sensitive
Oats
(Avena sativa)
Red clover
(Trifolium pratense)
Maize
(Zea mays)
Potato
(Solanum tuberosum)
Intermediate
Rye
(Secale cereale)
Wheat, common
(Triticum aesti vum)
Cabbage
(Brassica oleracea)
Resistant
Onion
(Allium cepa)
^ Sensitivity ratings were based on Van Haut and Stratmann (1967)
66
The most sensitive plant species may be injured by a two-hour exposure to
approximately 6.0 ppm of nitrogen dioxide under full sunlight conditions. On the other
hand, under cloudy conditions injury may occur through exposures to 2.5-3.0 ppm nitrogen
dioxide. It is important to note that in rural areas of Western Canada nitrogen dioxide
concentrations rarely exceed 0.10 ppm and, therefore, agricultural crops and plants are
rarely exposed to phytotoxic concentrations (Torn et al. 1987).
3.6.2.3 Foliar Effects of Ozone. Visible foliar injury as a result of ozone exposure
is almost always confined to green foliage of plants as opposed to fruits or floral
parts. The most common symptoms of VFI due to ozone as described by Hill et al. (1970)
are pigmented lesions, surface bleaching, bifacial necrosis, and chlorosis.
Leaves are most sensitive to ozone injury as they reach 65% to 95% of their
full size. Young leaves are generally resistant. The sensitivity of mature leaves
depends on the species. Conversely, young plants are more sensitive than mature plants
(Hill and Bennett 1970). Tingey et al. (1973b) reported a maximum sensitivity of soybean
to ozone induced foliar injury during the end of maximum leaf expansion when stomatal
resistance was low.
In fumigation experiments, ozone concentrations between 0.05 and 0.12 ppm for
two hours are usually required to injure the most sensitive species. Sensitive varieties
of alfalfa, spinach, clover, oats, sweet corn, and bean were injured by two hour
exposures at ozone concentrations of 0.10 to 0.12 ppm (Hill et al. 1970).
3.6.3 Growth and Yield Effects of Dry Deposition
Gaseous pollutants may cause either increases or decreases in growth and yield
with or without visible injury. Crop losses due to air pollutants have been reported
for over 30 years. In the United States, crop losses due to air pollution are estimated
to cost $1.8 billion per year (US); $1.7 billion is due to oxidants and $3.4 million is
due to sulphur dioxide (Stanford Research Institute 1981).
3.6.3.1 Effects of Sulphur Dioxide on Growth and Yield. Low concentrations of sulphur
dioxide can cause an increase in plant growth and yield in sulphur deficient soils.
Plants normally obtain sulphur in the form of sulphate absorbed from the soil, but when
soils are deficient, plants may compensate for foliar sulphur deficiency through
atmospheric sources. Most researchers have found that the increases in yield in the
presence of sulphur dioxide do not occur in plants grown in soils with sufficient sulphur
(Faller et al. 1970; Cowling and Lockyer 1978).
In addition, a plant may also utilize the sulphate in the soil derived from dry
and wet deposition. Jones et al. (1979) reported that atmospheric sulphur is a major
contributor to the needs of agronomic and horticultural crops as a plant nutrient in
South Carolina. Because soils are generally sulphur deficient and because commercial
fertilizers are quite expensive, atmospheric sulphur could be an important nutrient
source for farmers (Prince and Ross 1972).
Several studies have shown significant decreases in growth and yield due to
sulphur dioxide (Guderian 1977; Crittenden and Read 1978a, b; Heagle and Johnston 1979;
Davies 1980; Irving et al . 1982; Noggle and Jones 1982; and Heagle et al. 1983b).
67
The effects of sulphur dioxide from a nearby industrial source on the yield of
barley and alfalfa (at concentrations between 0.015 to 0.082 ppm (mean concentration over
three growing seasons) are shown in Table 16. At the highest concentration, barley
yield decreased by 34.9% and alfalfa yield by 30.3% when compared with the control
(Warteresiewicz 1979, cited in Godzik and Krupa 1982).
Guderian and Stratmann (1968) studied the effects of ambient sulphur dioxide on
various agricultural crops near an iron ore roasting plant. Their results indicated
that at an average sulphur dioxide concentration of 0.08 ppm, plant yields decreased by
as much as 9.1% for canola and 44.4% for winter wheat.
In similar studies, Maly (1974, cited in Godzik and Krupa 1982) reported
decreases in yield of 8.1 to 23.3% for various crops (Table 17). Of significance is
that although the pollutant concentrations in Maly's study were higher than those of
Guderian and Stratman (1968) yield reductions were of similar magnitude.
Response to a pollutant can differ among cultivars of the same species.
Laurence (1979) exposed seven varieties of wheat to various concentrations of sulphur
dioxide. The results of this experiment are shown in Table 18. At low SO2 concentra-
tions, Laurence showed that yields increased, but as concentrations increased the
cultivars responded with decreased yields. However, significant yield reductions were
not observed at commonly occurring dosage levels.
3.6.3.2 Effects of Nitrogen Oxide on Growth and Yield. Nitrogen dioxide in low concen-
trations can assume the role of a fertilizer and be a source of necessary nitrogen for
the plant. Investigators have reported increases in plant growth and yield with low
concentration nitrogen dioxide exposures. This fertilizer effect has been observed in
both nitrogen deficient and in nitrogen sufficient soils (Cowling and Koziol 1982).
Concentrations of 0.05 ppm nitrogen dioxide maintained continuously can cause small
reductions in growth and yield for sensitive agricultural species (Taylor et al. 1975).
The results of most of the studies conducted on nitrogen dioxide below 1.0 ppm
have shown inconclusive effects on growth and yield of agricultural plants. Spierings
(1971) studied the effects of nitrogen dioxide at a concentration of 0.25 ppm oh the
yield and growth characteristics of tomato over a 128 day growing season. He found that
there was a 22% decrease in fresh weight of the plants, a 12% decrease in average weight
of the fruits, and an 11% decrease in fruit number, as well as smaller leaves, petioles,
and stems. After 49 days at 0.25 ppm or at concentrations of 0.50 ppm after 10 days,
the plants grew tall and had thinner stems and smaller leaves. It is important to note
that the NO2 concentrations used in many of these experiments are seldom found under
ambient conditions.
3.6.3.3 Effects of Ozone on Growth and Yield. The phytotoxicity of ozone was firmly
established in 1957 (U.S. EPA 1978). Ozone has been shown to reduce growth and yield of
many agricultural species.
The lowest limit for injury resulting from ozone exposure follows long term
average concentrations of 0.02 to 0.05 ppm with periodic intermittant peaks, for most
species under general conditions (Guderian 1985). Results of experiments using acute
exposures of ozone are summarized in Table 19. The exposures in these experiments
varied in concentration from 0.05 to 1.0 ppm and an exposure time from one to 24 hours.
68
Table 16. Yields of two field crops grown in different concentrations
of sulphur dioxide.
Approximate Percent Yield:
SO2 Concentration Barley Grain Alfalfa Forage
(ppm)
.01 5
100.0
100.0
.029
98.0
99.2
.036
94.0
98.2
.038
92.2
100.4
.040
90.2
98.6
.047
85.9
88.0
.058
79.7
85.7
.060
76.3
82.0
.062
71 .8
76.3
.068
70.6
78.3
.079
64.7
70.0
.082
65.1
69.7
Source: Warteresiewicz (1979), cited by Godzik and Krupa (1982)
69
Table 17. Yield of various crops in field plots exposed to sulphur
dioxide.
Species: Decrease in Yield (%)
Oats, grain 12.2
Oats, straw 8.1
Clover 15.5
Cereals^ 20.0
(wheat, barley, rye & oats)
Potatoe 16.2
Flax (seed, fibre) 28.3. 23.8
Concentration: 1.26 ppm to 1.37 ppm (weekly averages)
Decrease in yield is relative to control
iData from a different growing season
Source: Maly (1974), cited by Godzik and Krupa (1982)
70
Table 18. Effects of sulphur dioxide on cuUivars of hard red spring
wheat (HRS) and soft white winter wheat (SWW).
Culti var
Concentration
(ppm)
Dry Weight After Exposure For:
30 h 78 h 100 h
Era (HRS)
0.0
0.089
0.112ab*
0.095ab
0.2
0.113
0.138a
0.114a
0.4
0.101
0.099ab
0.098ab
0.6
0.105
0.084b
0.076b
Waldron (HRS)
0.0
0.168
0.164
0.152ab
0.2
0.178
0.215a
0.173a
n A
0.142
0.165b
0.135ab
U . D
0.1 59
0.127b
0.108
Thatcher (HRS)
0.0
0.127
0.144
0.132ab
0.2
0.144
0.182
0.166a
0.125
0.143
0.127ab
U . 1 30
0.140
0.092b
Prelude (HRS)
0.0
0.1 65
0. 1 67
0.168
0.2
0.179
0.207
0.186
0.4
0.138
0.172
0.164
0.6
U . 1 / 1
n ICC
U . 1 DO
f\ Til
U. 1 22
Arrow (SWW)
0.0
0.146
0. 201 b
0.1 59ab
0.2
0.163
0.268a
0.178a
0 4
0.178
0.197b
0.1 38ab
n «
\J . 0
0.167
0.148b
0.117b
Ticonderoga (SWW)
0.0
0.156
0.154
0.151ab
0.2
0.147
0.172
0.174a
o!4
0.149
0.158
0.153ab
0.6
0.136
0.126
0.103b
Yorkstar (SWW)
0.0
0.162
0.145
0.157a
0.2
0.154
0.158
0.177a
0.4
0.161
0.145
0.145ab
0.6
0.141
0.120
0.095b
Means followed by the same letter are not significantly different
(P=0.05) based on Tukey's test for comparison of means. Absence of
letters indicates no significant difference. All comparisons are
made within one cultivar type and exposure period. Mean of 8
plants.
Adapted from the original table in Godzik and Krupa (1982)
Source: Laurence (1979)
Table 19. Effects of acute ozone exposure on growth and yield of
agricultural crops.
Plant Ozone Con- Exposure Plant Response^ Refer-
Species centration Time (h) Percent ence
(ppm) Reduction
Cucumber
cv. Ohio Mosaic
Grapevine
(Vitus labrusca)
cv. Ives
cv. Delaware
Pinto bean
Onion
cv. Sparan Era
1.0
1 .0
0.08
0.05
0.10
0.20
1.0
1.0
24
12
24
1
4
19, top dry wt
37, top dry wt
60, shoot growth
33, shoot growth
Significant re-
duction in leaf
growth
Significant re-
duction in leaf
growth
0, no effect
19, plant dry wt
49, plant dry wt
Potato
cv. Norland
Radish
cv. Cavalier
cv. Cherry Belle
Radish
1.0
0.25
0.25
0.40
4 0, tuber dry wt
4(3X ) 30, tuber dry wt
3 36, top dry wt
3 38, root dry wt
1.5(1X) 37, root dry wt
1 .5(2X) 63, root dry wt
1 .5(3X) 75, root dry wt
Snap bean
0.30
0.60
1.5 (2X)
1.5 (2X)
10, plant dry wt
12, pod dry wt
25, plant dry wt
41 , pod dry wt
continued.
72
Table 19. (Concluded)
Ozone Con- Exposure Plant Response^ Refer-
centration Time (h) Percent ence
(ppm) Reduction
Plant
Species
Soybean
0.30
to
0.45
Tall fescue
(Festuca arundinacea)
cv» Kentucky 3 0.30
1.5
2 (3X)
Threshold for re-
duction of shoot
growth
22, shoot dry wt
Tobacco
cv. Bel W3
Tomato
cv. Fireball
0.30
0.5
1 .0
0.5
1.0
48, chlorophyll 9
content
15, plant dry wt 10
(grown in moist soil)
20, plant dry wt
(grown in moist soil)
+15, plant dry wt
(grown in dry soil)
+25, plant dry wt
(grown in dry soil)
White clover
(Trifolium repens)
cv. Tillman
0.30
17, shoot dry wt
33, root dry wt
^ Responses marked with "+" are increases
References: 1. Ormrod et al . (1971)
2. Shertz et al . (1980)
3. Evans (1973)
4. Adedipe and Ormrod (1974)
5. Tingey et al . (1973a)
6. Blum and Heck (1980)
7. Heagle and Johnston (1979)
8. Kochhar et al . (1980), cited by Guderian (1985)
9. Adedipe et al . (1973)
10. Khatamian et al . (1973)
73
Various experiments on the effects of chronic exposure to ozone on agricultural
crops have been conducted and the results are summarized in Table 20.
3.6.4 Effects of Dry Deposition on Plant Reproduction
Dry deposition is suspected of having direct effects on plant reproductive
structures and processes. Unfortunately, aside from several studies on pollen germina-
tion, there has been little research on sulphur dioxide effects on plant reproduction.
A similar situation exists regarding the specifics of the reproductive effects of
nitrogen dioxide, although it has been known for several years that this gas causes
detrimental effects on reproductive structures of vegetation.
Pollen germination can be affected by sulphur dioxide exposure. Fumigations at
10 ppm for six days, with Swiss mountain pine and Scots pine pollen, caused a reduction
in germination and induced pollen tubes to burst when tests were run on a moist soil
medium (Dopp 1931). However, when these tests were run using a dry medium, no effects
were detected. Availability of moisture likely caused the formation of acids in the
presence of sulphur dioxide because without this moisture dry deposition did not appear
to cause a detrimental plant response, at least in these studies.
Decreases in the yield of fruits and seeds have been observed by several
investigators as a result of nitrogen dioxide fumigations (Taylor et al. 1975; Irving
et al. 1982; and Whitmore and Mansfield 1983). However, the specific numerical descrip-
tors of cause-effect relationships for these results are not available.
Ozone has been shown to cause decreases in grain or seed yield, number and
weight of fruit, and to delay fruit set. The aforementioned direct effects of ozone can
be found regardless of whether or not foliar injury (VFI) occurs (National Academy of
Sciences, U.S. 1977a; Bonte 1982; and Jacobson 1982).
Hydrogen sulphide at very low concentrations (0.07 ppm) caused a reduction in
catalase activity, seed germination, and size of the fruiting structure in brussel
sprouts (Dobrovolsky and Strikha 1970). Sprouts also showed signs of chlorosis. On the
basis of their results the authors concluded that in comparison with sulphur dioxide
toxicity, hydrogen sulphide was: ten times more toxic to seed germination; three times
more toxic to leaflet formation; two times more toxic with respect to sprout size; and
fifty times more inhibitory to catalase activity.
3.7 EFFECTS OF MIXTURES OF GASEOUS POLLUTANTS ON CROPS
The joint effects of pollutants on crops can be described as follows:
1. If the plant response equals the sum of the effects of the individual
pollutants, it is termed an additive response;
2. If the plant response to the combination of pollutants is greater than the
sum of the response to the individual pollutants, it is termed as a more
than additive response;
3. If the plant response to the combination of pollutants is lesser than the
sum of the response to the individual pollutants, it is termed as a less
than additive response.
74
Table 20. Effects of long-term controlled ozone exposures on growth,
yield, and foliar injury of various agricultural species.
Species Ozone
Cone,
(ppm)
Exposure
Hrs d-i/days
Alfalfa
Alfalfa
Bean,
pinto
Bean,
pinto
Bean,
pinto
Bean,
pinto
0.10
0.15
0.20
0.05
0.13
0.05
0.05
0.15
0.25
0.35
0.15
0.15
0.15
0.15
0.225
2/21 days
2/21 days
2/21 days
7/68 days
8/28 days
24 / 3-5 days
24/5 days
2/63 days
2/63 days
2/63 days
2/14 days
3/14 days
4/14 days
6/14 days
2/14 days
Plant Response Ref
(% Reduction or
Injury from Control)
16, top dry wt
26, top dry wt
39, top dry wt
30, shoot dry wt,
1st harvest
50, shoot dry wt,
2nd harvest
79, top dry wt
73, root fresh wt
70, height
50, leaf chlorosis
(fivefold increase in
lateral bud elongation)
33, plant dry wt
46, pod fresh wt
95, plant dry wt
99, pod fresh wt
97, plant dry wt
100, pod fresh wt
8, leaf dry wt
8, leaf dry wt
23, leaf dry wt
49, leaf dry wt
44, leaf dry wt
continued .
75
Table 20. (Continued)
Species Ozone
Cone .
(ppm)
Exposure
Hrs d~i/days
Plant Response
(% Reduction or
Injury from Control)
Ref
Bean,
pinto
Bean,
pinto
Beet
Crimson
clover
Corn,
sweet
cv. Golden
Fescue,
tall
Orchard
grass
0.225
0.30
0.30
0.06
0.20
0.03
0.20
0.35
0.05
0.10
0.09
0.09
Perennial 0.09
ryegrass
4/14 days
1 / 14 days
3/14 days
5 days/week
40 days
3/38 days
8/6 weeks
3/3 days/wk
until harvest
3/3 days/wk
until harvest
6/64 days
6/64 days
6 weeks
4/5 days/wk
5 weeks
4/5 days/wk
5 weeks
68, leaf dry wt
40, leaf dry wt
76, leaf dry wt
48, shoot dry wt
50, root dry wt
50, top dry wt
40, storage root dry wt
67, fibrous root dry wt
<10, dry wt
13, kernel dry wt 10
20, top dry wt
48, root dry wt
20, kernel dry wt
48, top dry wt
54, root dry wt
9, kernel dry wt
14, leaf injury
45, kernel dry wt
25, leaf injury
17, leaf dry wt 12
15, shoot dry wt
14 to 21. shoot 13
dry wt
14 to 21, shoot 13
dry wt
continued.
76
Table 20. (Continued).
Species Ozone
Cone .
(ppm)
Exposure
Hps d-i/days
Plant Response
(% Reduction or
Injury from Control)
Ref .
Potato 0.20
(2 seasons)
cv. Norland:
cv. Kennebec:
Potato
Radish
0.05
(>or =)
0.05
3 h (6X)
2 / week
326 to 533
total hours
two years
8/5 days/wk
5 weeks
14
30, tuber wt/19, tuber no.
20, tuber wt/21 , tuber no.
54, tuber wt/40, tuber no.
30, tuber wt/32, tuber no.
34 to 50. tuber 15
fresh wt
54, root fresh wt
16
Ryegrass, 0.09
Italian
8 /
6 weeks
36, dry wt
Soybean
Spinach
Soybean
Soybean
0.05
0.06
0.10
0.13
0.064
0.079
0.094
0.05
(>or=)
6/133 days
7 / day
37 days
9/55 days
465 /
growing season
3, seed yield
22, plant fresh wt
19, injury
18, fresh wt
37, fresh wt
69, fresh wt
31 , seed dry wt
45, seed dry wt
56, seed dry wt
28, seed wt
18
18
19
continued.
77
Table 20.
(Concluded) .
Species
Ozone
Cone .
(ppm)
Exposure
Hrs d~i/days
Plant Response Ref.
(% Reduction or
Injury from Control)
Tomato
0.20
2.5/3 days/wk
14 weeks
I. yield 10
32, top dry wt
I I , root dry wt
0.35
2.5/3 days/wk
45, yield; 72, top dry wt
59, root dry wt
8, tillering
Mheat
0.20
4/7 days
(anthesis)
30, yield 20
Wheat,
winter
0.10
0.13
7/54 days
7/54 days
16, seed dry wt 21
33, seed dry wt
References
1 .
Shinohara et al . (1974)
12.
Johnston et al. (1980)*
2.
Neely et al. (1977)*
13.
Horsman et al. (1980)*
3.
Manning et al . (1971a)
14.
Pell et al. (1980)*
4.
Engle and Gabelman (1966,1967)
15.
Heggestad (1973)
5.
Hoffman et al. (1973)
16.
Tingey et al. (1973a)
6.
Maas et al . (1973)
17.
Heagle et al. (1974)
7.
Manning (1978)*
18.
Heagle et al. (1979a)*
8.
Ogata and Maas (1973)
19.
Kress and Miller (1981)*
9.
Bennett and Runeckles (1977)
20.
Shannon and Mulchi (1974)
10.
Oshima (1973)
21 .
Heagle et al. (1979b)*
11 .
Heagle et al . (1972)
* References cited by Guderian (1985)
78
In sequential exposures, plants may become sensitized or tolerant to a pollutant
by a previous exposure to a different pollutant (Guderian 1985). Changes in injury type
may also occur in plants exposed to pollutant mixtures compared with single pollutants.
Plant responses to pollutant mixtures depend on many factors including components of the
mixture, temporal succession, and factors that influence a plant's response to single
pol lutants .
Sulphur dioxide from fuel combustion and ozone produced photochemical ly are the
two pollutants most frequently studied as mixtures. Because nitrogen dioxide is also
produced during combustion, it is also considered in certain studies on pollutant
mixtures. The potential for more than additive responses from mixtures of ozone,
sulphur dioxide and nitrogen dioxide is considered to be the most important pathway for
atmospheric interaction between plants and nitrogen dioxide (Taylor 1984).
Mechanisms for plant injury as a result of exposure to gaseous pollutant mixes
are not well understood, but it is assumed that the processes that govern plant responses
to single pollutants would hold (Guderian 1985). There does not seem to be any direct
correlation between visible symptoms and growth effects in plants exposed to gaseous
mixtures in contrast to the studies on single gaseous pollutants (Tingey et al. 1971a, b,
1973a, b; Mandl et al. 1973). It has been observed that growth of roots is inhibited
more than growth of other plant parts by exposures to gas mixtures, again in contrast to
effects documented for single gaseous pollutants (Ormrod 1984).
3.7.1 Combined Effects of Sulphur Dioxide and Ozone
Studies by Beckerson and Hofstra (1979a, b) on radish, cucumber, and soybean
exposed to sulphur dioxide and ozone both singly and in combination showed that stomatal
conductance was stimulated by SO2, inhibited by Oa, and inhibited to a greater
degree by the mixture. The concentration of each gas singly, or in the mixture, was
0.15 ppm.
Most of the studies on foliar injury due to sulphur dioxide and ozone have
indicated a less than additive effect. Few other studies, however, have shown more than
additive foliar effects of mixtures of the two gases.
In a number of studies, mixtures of sulphur dioxide and ozone have produced
more than additive effects on growth and yield at concentrations at or below a given
threshold. The yields of soybean, radish, and tobacco were found to respond additively
to a mixture of SO2 and Oa at concentrations of each between 0.05 and 0.10 ppm
(Tingey et al . 1971a, 1973c). Soybean root fresh weight was suppressed more than
additively by the pollutant mixture in this study. Other studies have also found the
responses of plants to mixtures of sulphur dioxide and ozone to be more than additive:
snapbean and tomato (Heggestad and Bennett 1981; Shew et al. 1982), or additive: potato,
fescue, soy beans (Flagler and Younger 1982; Foster et al . 1983; and Heagle et al.
1983b) .
3.7.2 Combined Effects of Sulphur Dioxide and Nitrogen Dioxide
Because ambient concentrations of nitrogen dioxide rarely approach the injury
thresholds for plants, its potential joint effects with other pollutants are a primary
concern. A majority of researchers have reported either additive or more than additive
79
effects in plants as a result of exposures to mixtures of sulphur dioxide and nitrogen
dioxide (Tingey et al. 1971b; Hill et al. 1974; Irving et al . 1982; Reinert and Sanders
1982; and others). A few investigators have also observed less than additive plant
responses (Thompson et al. 1980; Reinert and Sanders 1982; and Whitmore and Freer-Smith
1982).
Photosynthesis was shown to be initially stimulated in pea plants exposed to a
nitrogen dioxide/sulphur dioxide mixture; however, this effect was reversed leading to
inhibition within a short period of time (Bull and Mansfield 1974). Similarly, NO2-SO2
mixtures were found to decrease transpiration in bean plants at concentrations of 0.1 ppm
each even though individually these gases caused a stimulatory effect (Ashenden 1979).
Tingey et al. (1971b) observed more than additive injury on the adaxial surface
of leaves in plants exposed to a mixture of sulphur and nitrogen dioxides. This effect
differed greatly from the effects produced by either gas alone or from symptoms generally
produced by ozone. The visible injury threshold for the most sensitive agricultural
species with sulphur dioxide and nitrogen dioxide mixtures was between 0.05 and 0.10 ppm
of each gas (Tingey et al. 1971b).
Ashenden and Mansfield (1978) exposed four grass species to a mixture of sulphur
and nitrogen dioxide in long-term experiments using concentrations of 0.068 ppm each.
They observed a more than additive reduction, in the total dry weight of orchard grass,
Italian ryegrass, and timothy. In the case of Kentucky bluegrass, Ashenden and Mansfield
(1978) also found an additive growth reduction. Other studies on the joint effects of
sulphur dioxide and nitrogen dioxide on plants have shown: as a time series of growth,
more than additive, additive, or less than additive effects on the dry weight of grass
roots (Whitmore and Freer-Smith 1982); a more than additive reduction in seed production
in soybean (Irving et al. 1982); a less than additive reduction in root and shoot weight
in marigolds (Reinert and Sanders 1982); a general additive reduction of growth and
yield in desert plant species but in some instances, also less than additive effects
(Thompson et al . 1980) .
Research on the joint effects of sulphur dioxide and nitrogen dioxide on plant
reproduction is limited. Masaru et al. (1976) have shown that pollen tube growth can be
reduced more than additively in lilies exposed to S02:N02 at concentrations of
0.24:0.12 ppm for 30 to 60 minutes.
From the literature review, it is evident that the physiological effects of
mixtures of sulphur dioxide and nitrogen dioxide on plants is not well understood. What
is known indicates that effects are species specific and often quite different from the
responses to individual pollutant exposures.
3.7.3 Combined Effects of Nitrogen Dioxide and Ozone
Very few studies have been conducted on the effects of mixtures of nitrogen
dioxide and ozone on plants. More than additive, additive, and less than additive
interactions have been documented with respect to foliar injury (Torn et al. 1987). The
temporal sequencing of exposures to NO2 and Oa has been shown to be important. In
one study, increased sensitivity of plants to ozone was observed as a result of previous
exposure to nitrogen dioxide. This resulted in reduced growth and yield (Runeckles
et al. 1978). While not immediately relevant to crops, studies on growth and biomass in
80
trees exposed to mixtures of nitrogen dioxide and ozone have produced a variety of
results ranging from more than additive growth suppression (height growth in Virginia
and Loblolly pine) to less than additive suppression of root accumulation in sweet gum
and total dry weight in white ash (Kress and Skelly 1982).
3.7.4 Combined Effects of Sulphur Dioxide, Nitrogen Dioxide, and Ozone
Studies on the joint effects of sulphur dioxide, nitrogen dioxide, and ozone
are in the initial phases and little definitive documentation exists at this time (Torn
et al. 1987).
Foliar injury similar to that observed with ozone alone has been observed when
plants were exposed to the combination of the three pollutants (Torn et al . 1987), The
effects on growth and yield of sulphur dioxide, nitrogen dioxide, and ozone have not
been thoroughly studied. Researchers have found that in nearly every instance, exposure
to the three pollutants caused a greater loss in plant growth and yield than exposure to
a single or a mixture of two pollutants. Studies conducted thus far are important
because they have shown that growth and yield responses to the three pollutant mixture
occur in the nitrogen dioxide concentration range of 0.05 to 0.30 ppm, i.e., within the
ambient concentrations of nitrogen dioxide at certain locations (Torn et al. 1987). The
decrease in growth and yield caused jointly by nitrogen dioxide with sulphur dioxide
and/or ozone varied from 5% to 20% at concentrations of nitrogen dioxide that cause
little or no injury when that pollutant was used singly (Reinert 1984).
3.8 COMBINED EFFECTS OF DRY AND WET DEPOSITION
The joint effects of dry and wet deposition are considered to be significant
since both processes contribute to the pollutant burden. Research to date suggests that
more than additive or additive effects on plant processes and foliar injury can occur as
a result of exposure to both types of deposition (Torn et al . 1987).
Experiments using various ozone concentrations and simulated acidic precipi-
tation with differing pH values produced an additive response in foliar injury and
reduction of chlorophyll in young leaves of radish (Shriner 1983). Older leaves in the
same experiment exhibited a more than additive foliar injury response. Similar results
were reported by Shriner (1983) for simulated acidic precipitation and sulphur dioxide.
On the contrary, experiments conducted by Norby and Luxmoore (1983) on soybean exposed
to simulated acidic rain and a gaseous mixture of ozone and sulphur dioxide showed no
foliar injury effects. The effects of of wet and dry deposition on growth and yield
have been studied by Shriner (1978, 1983) and Irving and Miller (1981). These studies
showed either additive or more than additive plant responses to combinations of wet and
dry deposition, depending on the species studied.
3.9 EFFECTS OF ACIDIC DEPOSITION ON PLANT-SOIL INTERACTIONS
Ihe effects of wet and dry deposition on soils have been fully discussed in an
earlier section and will only be treated here as they relate directly to plants and
agricultural practices.
Short-term impacts of acidic rain or gaseous pollutants on agricultural soils
will be small on intensively managed systems (Coleman 1983; McFee 1983; Mortvedt 1983;
81
and Torn et al. 1987). At current or projected levels of ambient acidity, N and S
deposited in acidic rain will act as fertilizer supplements rather than as toxins on
soils of all degrees of management (Jones and Suarez 1979; Sandhu et al. 1980). The
greater potential for toxicity relates to the free hydrogen concentration of acidic rain.
However, current agricultural practices have a much greater effect on soil pH than does
atmospheric deposition. Estimates are that the H"*" flux from heavily acidified rain
would be only 1% of the total flux available from nitrogen fertilizers (Plocher et al.
1985).
On less intensively managed lands, acidic precipitation could have a significant
effect on soil quality and the overall fertility of the soil. While acidification is
prevented or managed on agricultural soils, it is essentially irreversible in unculti-
vated soils without excessive expenditure (McFee 1980). Generalized responses of the
soil environment to natural or anthropogenic changes in soil pH are summarized in
Figure 1 (Brady 1974). The most likely changes in soil characteristics that could
result from acidic deposition are: increase in the acidity of soil solution; increase
in exchangeable aluminum and other heavy metals; and a change in the composition of the
exchangeable ion complex with a concomitant decrease in base saturation capacity
(Russell 1973; Agrawal et al . 1985).
Very few investigators have addressed the potential for acidic deposition
induced changes in plant growth, in the context of the interactions of the plant with
the soil. A possible exception is symbiotic systems in roots (Torn et al . 1987).
3.9.1 Effects of an Acidified Soil Environment on Plants
The threshold for direct toxicity to plants from the acidity of soil solution
is pH 3.0 (Russell 1973). Before the soil becomes this acidic, related changes in the
soil will render the soil unsuitable for most crops; these secondary effects are the
primary route by which soil acidity is harmful (Russell 1973). High aluminum concentra-
tion is the most common cause of crop failure on acidic soils (Russell 1973). Aluminum
can harm plants in two ways. Aqueous aluminum in free spaces of the root surface may
inhibit root uptake of phosphates, and sugar phosphorylation may be inhibited by inter-
cellular aluminum. Manganese toxicity can also result from increased acidity (Russell
1973). Tables 21 and 22 indicate the concentrations of various heavy metals toxic to
plants and plant sensitivity to such metals under acidity induced soil conditions.
Overall soil acidity also has an effect on the ability to grow certain crops.
Table 23 shows the recommended crops for Great Britain under varying soil pH. Table 24
shows the effects of reduced soil pH on the yield of barley and alfalfa.
3.9.2 Effects of Altered Soil Acidity on Soil Organism-Plant Interactions
The pH of the soil influences the success of soil-borne organisms which can
either be beneficial or harmful to plants. Some soil organisms are at an advantage in
acidic soils, whereas others are inhibited. Therefore, the net effect of acidic
deposition or gaseous pollutant exposure on plant health will encompass four factors:
(1) the deleterious effects on commensals or symbionts; (2) the stimulation or inhibition
of pests; (3) the stimulation or inhibition of plant health; and (4) the effects of
altered plant biochemistry on plant-organism interactions.
82
PH
Fungi
8
Bacteria and octinomycetes;
N
Ca and Mg
Fe,Mn, Zn,Cu,Co
Mo
B
Figure 1. Relationship between soil pH and activity of microorganisms
and availability of plant nutrients.
Source: Brady (1974)
83
Table 21. Toxic concentration of copper, nickel, or zinc in leaf
tissue.
opec 1 es
LU
(ppm)
m
7n
Spring barley
19
12
210
Ryegrass
21
14
221
Lettuce
21
Canola
16
Wheat
18
Source: Davis and Beckett (1978)
Table 22. Plant sensitivity to acid-induced changes in the soil
environment.
Aluminum tolerance Oats » potatoes » beets
Manganese tolerance Oats » beets » potatoes
Calcium demand Beets = potatoes » oats
Source: Russell (1973)
84
Table 23. Recommended crops for soils with varying acidity in
Great Britain.
Soil Acidity
Crops Recommended
Neutral to low acidity
Alfalfa
Barley
Sugar beet
Medium acidity
Peas
Red clover
Wheat
High acidity
Oats
Rye
White clover
Source: Russell (1973)
Table 24. Effect of drop of 0.1 unit in soil pH on barley and
alfalfa yield.
Initial
Crop
pH Range
Reduction in Yield
Barley
5.2 - 5.5
161 kg/ha
Alfalfa
5.5 - 6.0
448 kg/ha
Source: Sandhu et al. (1980)
85
3.9.2.1 Effects of Acidic Deposition on Plant-Microbe Interactions. The effects of
acidic precipitation and gaseous dry deposition on the life cycles of pathogens, on
plant vulnerability to pathogens, and on plants as hosts for beneficial organisms is a
topic of intensive study in agriculture. Table 25 provides a comprehensive listing of
experiments conducted with the major gaseous pollutants and acid precipitation on
agricultural crops and their pathogens.
3.9.2.2 Effects on the Plant as Host Organism. Certain characteristics of the
host-pathogen relationship appear to be sensitive indicators of the overall stress on
the plant due to gaseous and aqueous pollutants (Shriner 1980). Pollutants may alter
the primary metabolites, affect the digestibility of the host and the feeding behaviour
of insects (Hughes 1983). For example, Mexican bean beetles developed more slowly and
were less fecund when feeding on plants fumigated with hydrogen fluoride (Hughes 1983).
Pollutants may reduce the ability of the plant to produce defensive chemicals. Produc-
tion of protective chemicals may also be reduced or increased by insect injury to the
plant (Hughes 1983; Schultz and Baldwin 1982, respectively). Therefore, plant injury
symptoms must be carefully examined before assuming a cause-effect relationship as a
result of acidic deposition.
3.9.2.3 Effects on Viruses, Fungi, and Bacteria. Stimulation or inhibition of growth
and reproduction due to acidic precipitation has been shown to vary widely for soil
bacteria, yeasts, and fungi. Bacteria are the least resistant to acidity (activity
below pH 5.6 is reduced and is almost zero at pH 4.0 (Brady 1974, Figure 7), while fungi
are the most tolerant (Shriner 1978). Table 26 summarizes information on the stages in
the life cycles of fungi with the potential for the greatest interference by acidic
pollutants,
Hughes and Laurence (1983) reported that viruses are more successful on plants
exposed to air pollution. Conversely, in the studies conducted by Laurence (1981) and
Hughes and Laurence (1983), viral infection provided protection to plants exposed subse-
quently to ozone. The incidence and severity of diseases caused by obligate fungal
parasites can also be reduced by exposure to ozone (Hughes 1983). Hughes also found that
multiple exposures to sub-acute concentrations of ozone increased the success of powdery
mildew on barley plants. Air pollution stress may increase the incidence and severity
of diseases induced by non-obligate fungal parasites (Laurence 1981). This is considered
to be important since these are numerous, widely distributed facultative parasites often
associated with important agricultural crops (Laurence 1981).
3.9.2.4 Effects on Insect-Plant Relationships. In discussing research on pollutant-
plant-insect interactions, Hughes (1983) noted that sulphur dioxide stimulated growth
and reproduction of the milkweed bug, as did carbon monoxide and nitric oxide. Feir
(1978) concluded that feeding insects are relatively unaffected by contact with gaseous
pollutants such as ozone, but are significantly affected by water-soluble pollutants
such as acidic sulphate aerosols.
Beetles were found to be more fecund and grew to a larger size in fields of
soybean exposed to sulphur dioxide (Hughes et al. 1981, 1982; Hughes 1983). Hughes
86
Table 25. Effect of pollutants on plant-pathogen interactions.
Plant/Pathogen Exposure^ Effect on Pathogen Effect on Ref.
Induced Disease Pollutant
Injury on
the Plant
OZONE
Barley/
Ervsiphe graminis
Wheat/ S
Puccinia graminis
Oats/P^ coronata S
Oats/P_^ coronata S
Wheat/P_^ graminis A
Wheat/P^ graminis A
Corn/ S
Helminthosporum mavdis
Race T
Geranium/Botrytis A
cinerea
Geranium/B. cinerea A
Broad bean/B^ cinerea -
Potato/B^ cinerea
Geranium flowers/ A
B^ cinerea
Geranium leaves/ A
B^ cinerea
Poinsettia/ A
B^ cinerea
Pinto bean/Root A
inhabiting fungi
Cabbage/Fusarium S
oxysporium
Rose/Diplocarpon S
rosae
Reduced infection from
exposed spores, colony size
reduced. Multiple exposure
caused increases in colony
size.
Reduced sporulation.
Reduced sporulation.
Reduced growth of uredia.
Decreased growth of hyphae.
Decreased number of spores.
Reduced infection.
a) 18 pphm increased colony
size
b) 12 pphm increased number
of spores
Reduced sporulation.
Reduced infection by exposed
spores.
Flocculent material produced
Increased disease development.
Predisposition to infection.
Reduced disease development.
Increased disease development.
No effect on disease
development.
Increased number of fungal
colonies .
Decreased nodulation.
Decreased disease development
slightly.
Reduced disease development.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity,
8
9
10
n
12
9
13
14
4
continued .
87
Table 25 (Continued) .
Plant/Pathogen Exposure^ Effect on Pathogen Effect on Ref.
Induced Disease Pollutant
Injury on
the Plant
OZONE
Tobacco/Tobacco F
mosaic virus
Tobacco, Pinto bean/ A
Tobacco mosaic virus
Pinto bean/bean A
common mosaic virus
Pinto bean/alfalfa A
mosaic virus, tobacco
ringspot virus,
tobacco mosaic virus,
tomato ringspot virus
Tobacco/Tobacco etch A
vi rus
Tobacco/Tobacco A
streak virus
Soybean/Rhizobium A
japonicum
Alfalf a/Xanthomonas A
alfalfa
Kidney bean/Pseudo A
monas phaseolicola
Soybean/Pseudomonas A
sp.
Soybean/P_^ glycinea A
Wild strawberry/ A
Xanthomonas f ragaiae S
Root growth and nodulation
reduced.
Reduced disease development.
Increased and modified
Hypersensitive reaction (HR)
( Pre-exposure inoculation).
No HR (Post-exposure
inoculation) .
Reduced disease incidence.
Reduced disease incidence.
No effect.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity.
Increased Oa
sensitivity.
Reduced Oa
sensitivity.
Reduced Oa
sensitivity
in halo.
Inoculation
24 h before
exposure
prevented Oa
injury.
No effect.
No effect.
15
16
12
13
17
18
19.20
21
22
23
24
25
SULPHUR DIOXIDE
Wheat/Puccinia
graminis
Corn/Helminthosporium
maydis
Bean/southern bean
mosaic virus
Reduced disease
development.
Reduced disease
development.
Increased virus titer.
No effect. 26
26
Increased 27
sulfur uptake.
continued.
88
Table 25 (Continued).
Plant/Pathogen Exposure^ Effect on Pathogen Effect on Ref.
Induced Disease Pollutant
Injury on
the Plant
SULPHUR DIOXIDE
Corn/maize dwarf
mosaic virus
Tomato/tobacco mosaic
virus
Corn/Corvnebacterium
nebraskense
Soybean/Mex. bean
beetle
Increased virus titer. No effect. 27
Increased symptom severity.
No effect. No effect. 27
Reduced and delayed disease No effect. 28
development.
Increased beetle fecundity. - 29
ACIDIC PRECIPITATION
Corn/He 1 mi nthospori um A
maydis (N cytoplasm)
Corn/H_^ maydis A
(N cytoplasm)
Corn/H^ maydis
Kidney bean/Uromyces A
phaseoli
Kidney bean/ A
Pseudomonas
phaseolicola
Kidney bean/root knot A
nematode
Soybean and kidney A
bean/Rhizobium sp.
Phaseolus vulgaris/
Meloidoqyne hapla
Phaseolus vulgaris/
Uromyces phaseoli
Phaseolus vulgaris/
Pseudomonas phaseol icola
Increased disease development.
No effect.
No effect.
Decreased disease development.
Post-exposure inoculation;
increased disease development.
Decreased disease development.
Decreased nodulation.
No effect.
Inhibited growth.
Inhibited growth.
No effect
No effect.
No effect.
Increased
injury.
30
30
31
30,32
30,32
30,32
32
31
31
31
^ F = Field exposure
S = Sub-acute exposure
A = Exposure causing acute injury
F(A) = Field exposure with acute injury
- = Information not available
continued .
89
Table 25 (Concluded).
References 1-28, and 30 cited by Laurence (1981)
References: 1. Heagle and Strickland (1972)
2. Heagle (1975)
3. Heagle (1970)
4. Heagle and Key (1973)
5. Treshow et al. (1967)
6. Heagle (1977)
7. Krause and Weidensaul (1978a)
8. Krause and Weidensaul (1978b)
9. Manning et al. (1972)
10. Manning et al . (1969)
11 . Manning et al. (1970b)
12. Manning et al. (1970a)
13. Manning et al . (1971a)
14. Manning et al. (1971b)
15. Bisessar and Temple (1977)
16. Brennan (1975)
17. Moyer and Smith (1975)
18. Reinert and Gooding, Jr. (1978)
19. Blum and Tingey (1977)
20. Tingey and Blum (1973)
21 . Howell and Graham (1977)
22. Kerr and Reinert (1968)
23. Pell et al. (1977)
24. Laurence and Wood (1978a)
25. Laurence and Wood (1978b)
26. Laurence et al. (1979a)
27. Laurence et al . (1979b)
28. Laurence (unpublished data)
29. Hughes et al. (1983)
30. Shriner (1977)
31. Shriner (1980)
Adapted from Laurence (1981)
90
Table 26. Simulated acid rain-fungal life cycle interaction (Torn et
al. 1987).
Stage
Process
Effect of Low pH
Spore dissemination
Dissemination by
water splash
Germination
inhibited
Spore on tissue
Spore germinates
and grows prior
to penetration
Growth stimulated
or inhibited
Penetration
Penetration
through cuticle
stomata or wounded
tissue
Wounded tissue
increase, stomata
close, cuticle eroded
Colonization
Pathogen dependent
on host metabolism
Changes in primary
and secondary
metabolites
References: 1. Shriner and Cowling (1980)
2. Laurence et al . (1983)
91
et al. (1982) also found that female beetles preferred to feed on sulphur dioxide
fumigated young plants first, unfumigated mature plants next, and unfumigated young
plants last. He suggested that sulphur dioxide may induce physiological changes in
young plants similar to those which occur in older plants and as a result, a shift in
the preference of the beetles. The overall result of sulphur dioxide fumigation may be
an extension in the period of vulnerability of plants to predation because under normal
conditions, the insects feed primarily on older mature plants. Similar types of insect-
plant relationships may also occur with other plant species and pollutants.
92
3.10 ACIDIC DEPOSITION EFFECTS ON AGRICULTURE: LITERATURE CITED
Adams, CM. and T.C. Hutchinson. 1984. A comparison of the ability of leaf surfaces of
3 species to neutralize acidic rain drops. New Phytologist 97: 463-478.
Adedipe, N.O. and D.P. Ormrod. 1974. Ozone-induced growth suppression in radish plants
in relation to pre- and post-fumigation temperatures. Zeitschrift
Pf lanzenphysiologia 71: 281-287.
Adedipe, N.O., R.A. Fletcher, and D.P. Ormrod. 1973. Ozone lesions in relation to sen-
escence of attached and detached leaves of tobacco. Atmospheric Environment 7:
357-361 .
Agrawal, M., P.K. Nandi, and D.N. Rao. 1985. Effects of sulphur dioxide fumigation on
soil system and growth behaviour of Vicia f aba plants. Plant and Soil 86: 69-78.
Alberta Agriculture. 1982. 1981 Census of Agriculture for Alberta. Edmonton, Alberta.
217 pp.
Amthor, J.S. 1984. Does acid rain directly influence plant growth? Some comments and
observations. Environmental Pollution, Series A 36(1): 1-6.
Amthor, J.S. and F.H. Bormann. 1983. Productivity of perennial ryegrass as a function
of precipitation acidity (Lolium perenne) . Environmental Pollution, Series A
32(2): 137-145.
Ashenden, T.W. 1979. The effects of long-term exposures to sulfur dioxide and nitrogen
dioxide pollution on the growth of Dactyl i s qlomerata L. and Poa pratensis L.
Environmental Pollution 18: 249-258.
Ashenden, T.W. and T.A. Mansfield. 1978. Extreme pollution sensitivity of grasses when
SO2 and NO2 are present in the atmosphere together. Nature 273: 142-143.
Baddeley, M.S. and B.W. Ferry. 1973. Air Pollution and Lichens. London: Athlone Press.
313 pp.
Barrett, T.W. and H.M. Benedict. 1970. Sulfur dioxide. In: Recognition of Air Pollution
Injury to Vegetation: A Pictorial Atlas, eds. J.S. Jacobson and A.C. Hill.
Pittsburgh: Air Pollution Control Association, pp. C1-C17.
Beckerson, D.W. and G. Hofstra. 1979a. Stomatal responses of white bean to ozone and
sulfur dioxide singly or in combination. Atmospheric Environment 13: 533-535.
Beckerson, D.W. and G. Hofstra. 1979b. Response of leaf diffusive resistance of radish,
cucumber and soybean to O3 and SO2 singly or in combination. Atmospheric
Environment 13: 1263-1268.
Bennett, J. P. and V.C. Runeckles. 1977. Effects of low levels of ozone on plant compe-
tition. Journal of Applied Ecology T4:877-880.
Bisessar S. and P.J. Temple. 1977. Reduced ozone injury on virus infected tobacco in
the field. Plant Disease Reporter 61: 961-963.
Black, V.J. 1982. Effects of sulphur dioxide on physiological processes in plants. In:
Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterworth Scientific, pp. 67-91.
Blum, U. and D.T. Tingey. 1977. A study of the potential ways in which ozone could
reduce root growth and nodulation of soybean. Atmospheric Environment 11:
737-739.
Blum, U. and W.W. Heck. 1980. Effects of acute ozone exposure on snap bean at various
stages of its life cycle. Environmental and Experimental Botany 20: 73-85.
Bonte, J. 1982. Effects of air pollutants on flowering and fruiting. In: Effects of
Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H. Unsworth and
D.P. Ormrod. London: Butterworth Scientific, pp. 207-223.
93
Brady, N.C. 1974. The Nature and Properties of Soils. New York: Macmillan Publishing.
639 pp.
Brennan, E. 1975. On exclusion as the mechanism of ozone resistance of virus-infected
plants. Phytopathology 65: 1054-1055.
Brewer, P.F. and A.S. Heagle. 1983. Interactions between Glomus qeosporum and exposure
of soybeans to ozone or simulated acid rain in the field. Phytopathology
73(7): 1035-1040.
Bruton, V.C. 1974. Environmental influence on the growth of Arabidopsis thai iana.
Raleigh: North Carolina State University. 105 pp. Ph.D. Thesis.
Bull, J.N. and T.A. Mansfield. 1974. Photosynthesis in leaves exposed to sulfur dioxide
and nitrogen dioxide. Nature 250: 443-444.
Cohen, C.J., L.C. Grothaus, and S.C. Perrigan. 1982. Effects of simulated sulfuric and
sulf uric-nitric acid rain on crop plants: results of 1980 crop survey. Special
Report 670, November. Corvallis, Oregon: Agricultural Experiment Station, Oregon
State University. 92 pp.
Coleman, D.C. 1983. The impacts of acid deposition on soil biota and C (Carbon) cycling.
Environmental and Experimental Botany 23(3): 22.
Cowling, D.W. and D.R. Lockyer. 1978. The effects of SO2 on Lol ium perenne L. grown
at different levels of sulphur and nitrogen nutrition. Journal of Experimental
Botany 29: 257-265.
Cowling, D.W. and M.J. Koziol. 1982. Mineral nutrition and plant response to air pol-
lutants. In: Effects of Gaseous Air Pollution in Agriculture and Horticulture,
eds. M.H. Unsworth and D.P. Ormrod. London: Butterworth Scientific. pp.
349-375.
Cox, R.M. 1983. Sensitivity of forest plant reproduction to long range transported air
pollutants: in vitro sensitivity of pollen to simulated acid rain. New
Phytologist 95: 269-276.
Cox, R.M. 1982. Determination of the sensitivity of the pollination processes of dif-
ferent forest flora species to simulated acid rain. Toronto, Ontario: Insti-
tute of Environmental Studies, University of Toronto. 41 pp.
Crafts, A.S. 1961. The chemistry and mode of action of herbicides. New York: Inter-
science Publishers.
Craker, L.E. and D. Bernstein. 1984. Buffering of acid rain by leaf tissue of selected
crop plants. Environmental Pollution, Series A 36(4): 375-381.
Crittenden, P.O. and D.J. Read. 1978a. The effects of air pollution on plant growth
with special reference to sulfur dioxide. Part I: Introduction and chamber
conditions. New Phytologist 80: 33-48.
Crittenden, P.D. and D.J. Read. 1978b. The effects of air pollution on plant growth
with special reference to sulfur dioxide. Part II: Growth studies with Lol ium
perenne L. New Phytologist 80: 49-62.
Davies, T. 1980. Grasses more sensitive to SO2 pollution in conditions of low irradi-
ance and short days. Nature 284: 483-485.
Davis, R.D. and P.H.T. Beckett. 1978. Upper critical levels of toxic elements in plants
II. Critical levels of copper in young barley, wheat, rape, lettuce, and
ryegrass, and of nickel, and zinc in young barley and ryegrass. New Phytologist
80: 23-32.
De Koning, H.W. and Z. Jegier. 1968. A study on the effects of ozone and sulfur dioxide
on the photosynthesis and respiration of Euglena graci 1 is . Atmospheric Envi-
ronment 2: 321-326.
94
Dobrovolsky, I. A. and E.A. Strikha. 1970. Study of the phytotoxicity of certain com-
ponents of industrial air pollution. Lawrence Berkeley Laboratory. (Translated
by Oleh Weres, Ph.D. ) .
Dopp, W. 1931. Bericht der Deutschen Botanischen Gesellschaft 49: 172-221.
Dreisinger, B.R. and P.C. McGovern. 1970. Monitoring SO2 and correlating its effects on
crops and forestry in the Sudbury area. I_n: Proceedings of the Conference on
Impact of Air Pollution on Vegetation, ed. S.N. Linzon. 1970 April 7-9;
Toronto, Ontario; pp. 12-28.
Engle, R.L. and W.H. Gabelman. 1967. The effects of low levels of ozone on pinto beans,
Phaseolus vulgaris L.. Ijn: Proceedings of American Society of Horticultural
Science 91 : 304-309.
Engle R.L. and W.H. Gabelman. 1966. Inheritance and mechanism of resistance to ozone
damage to onion, Allium cepa L. Proceedings of the American Society of Horti-
cultural Science 89: 423-430.
Evans, L.S. 1984. Botanical Aspects of Precipitation. Botanical Revue 50: 449-490.
Evans, L.S. 1973. Bean leaf growth response to moderate ozone levels. Environmental
Pollution 4: 17-26.
Evans, L.S. and T.M. Curry. 1979. Differential responses of plant foliage to simulated
acid rain. American Journal of Botany 66(8): 953-962.
Evans, L.S., K.F. Lewin, and M.J. Patti. 1984. Effects of simulated acidic rain on
yields of field-grown soybeans. New Phytologist 96(2): 207-213.
Evans, L.S., K.F. Lewin, M.J. Patti, and E.A. Cunningham. 1983. Productivity of field-
grown soybeans exposed to simulated acidic rain. New Phytologist 93(3): 377-388.
Evans, L.S., K.F. Lewin, and E.A. Cunningham. 1982a. Effects of simulated acidic rain
on yields of field-grown radishes and garden beets. Agriculture and Environment
7: 285-289.
Evans, L.S., K.F. Lewin, E. A. Cunningham, and M. J. Patti. 1982b. Effects of simulated
acidic rain on yields of field-grown crops. New Phytologist 91: 429-441.
Evans, L.S., N.F. Gmur, and D. Mancini. 1982c. Effects of simulated acidic rain on yields
of Raphanus sati vus . Lactuca sativa, Triticum aesti vum and Medicago sati va.
Environmental and Experimental Botany 22: 445-453.
Evans, L.S., G.R. Hendrey, G.J. Stensland, D.W. Johnson, and A.J. Francis. 1981a. Acidic
precipitation: considerations for an air quality standard. Water, Air, and
Soil Pollution 16: 469-509.
Evans, L.S., T.M. Curry, and K.F. Lewin. 1981b. Responses of leaves of Phaseolus vulgaris
L. to simulated acid rain. New Phytologist 88: 403-420.
Evans, L.S., C. A. Conway and K.F. Lewin. 1980. Yield responses of field-grown soy beans
exposed to simulated acid rain. I_n: Ecological Impact of Acid Precipitation,
eds. D. Drablos and A. Tollan; SNSF Project, As-NLH, Norway; pp. 162-163.
Evans, L.S., N.F. Gmur, and F. D'Acosta. 1977. Leaf surface and histological perturba-
tions of leaves of Phaseolus vulgaris and Hel ianthus annuus after exposure to
simulated acid rain. American Journal of Botany 64(7): 903-913.
Faller, N., K. Herwig, and H. Kuhn. 1970. The uptake of sulfur-35 dioxide from the air
- Part 1: Effect on plant yield. Plant and Soil 33: 177-191.
Feir, D. 1978. Effects of air pollution on insect growth and reproduction. Physiologist
21: 36.
Ferenbaugh, R.W. 1976. Effects of simulated acid rain on Phaseolus vulgaris L. (Faba-
ceae). American Journal of Botany 63(3): 283-288.
95
Flagler, R.B. and V.B. Younger. 1982. Ozone and sulfur dioxide effects on tall fescue:
I. Growth and yield responses. Journal of Environmental Quality 11: 386-389.
Forsline, P.L, R.C. Musselman, R.J. Dee, and W.J. Kender. 1983a. Effects of acid rain
on grapevines. American Journal of Enology and Viticulture 34(1): 17-22.
Forsline, P.L., R.C. Musselman, W.J. Kender and R.J. Dee. 1983b. Effects of acid rain on
apple tree productivity and fruit quality. Journal of the American Society of
Horticultural Science 108: 70-74.
Foster, K.W., H. Timm, C.K. Labanauskas, and R.J. Oshima. 1983. Effects of ozone and
sulfur dioxide on tuber yield and quality of potatoes. Journal of Environmental
Quality 12: 75-79.
Fujiwara, T. 1975. Studies on the development of injury symptoms caused by sulphur
dioxide at low level concentrations in plants and the diagnosis of injury.
Bulletin of the Agricultural Laboratory CRIEPI, No. 74001.
Garcia, R.L. and J.J. Hanway. 1976. Foliar fertilization of soybeans during the seed-
filling period. Agronomy Journal 68(4): 653-657.
Gilbert, O.L. 1968. Biological indicators of air pollution. Newcastle, England: Univer-
sity of Newcastle-Upon-Tyne. Ph.D. Thesis.
Godzik, S. and S.V. Krupa. 1982. Effects of sulphur dioxide on the growth and yield of
agricultural and horticultural crops, in: Effects of Gaseous Air Pollution in
Agriculture and Horticulture, eds. M.H. Unsworth and D.P. Ormrod. London:
Butterworth Scientific, pp. 247-265.
Guderian, R. 1985. Air pollution by photochemical oxidants. New York: Springer-Verlag.
346 pp.
Guderian, R. 1977. Air Pollution - Phytotoxicity of Acidic Gases and its Significance
in Air Pollution Control. New York: Springer-Verlag. 127 pp.
Guderian, R. and H. Strattmann. 1968. Frei landversuche zur Ermittlung von Schwef eldiox-
idwirkungen auf die Vegetation. III. Teil Grenzwerte schadlicher SO2-
Immissionen fur Obstund Forstkulturen sowie fur landwirtschaftliche und
gartnerische Pf lanzenarten. Forschungsberichte des Landes NordrheinWestfalen.
114 pp.
Harcourt, S.A. and J.F. Farrar. 1980. Some effects of simulated acid rain on the growth
of barley and radish. Environmental Pollution 22: 69-73.
Heagle, A.S. 1977. Effect of ozone on parasitism of corn by Helminthosporium maydis.
Phytopathology 67: 616-618.
Heagle, A.S. 1975. Responses of three obligate parasites to ozone. Environmental Pol-
lution 9: 91-95.
Heagle, A.S. 1970. Effect of low-level ozone fumigations on crown rust of oats. Phyto-
pathology 60: 252-254.
Heagle, A.S. and J.W. Johnston. 1979. Variable responses of soybeans to mixtures of
ozone and sulfur dioxide. Journal of the Air Pollution Control Association 29:
729-732.
Heagle, A.S. and L.W. Key. 1973. Effect of ozone on the wheat stem rust fungus. Phyto-
pathology 63: 397-400.
Heagle, A.S. and A. Strickland. 1972. Reaction of Erysiphe graminis f . sp. hordei to low
levels of ozone. Phytopathology 63: 1144-1148.
Heagle, A.S., W.W. Heck, J.O. Rawlings, and R.B. Philbeck. 1983. Effects of chronic
doses of ozone and sulfur dioxide on injury and yield of soybeans in open-top
field chambers. Crop Science 23: 1184-1191.
96
Heagle, A.S., R.B. Philbeck, H.H. Rogers, and M.B. Letchworth. 1979a. Dispensing and
monitoring ozone in open-top field chambers for plant-effects studies. Phyto-
pathology 69: 15-20.
Heagle, A.S., S. Spencer, and M.B. Letchworth. 1979b. Yield response of winter wheat to
chronic doses of ozone. Canadian Journal of Botany 57: 1999-2005.
Heagle, A.S., D.E. Body, and G.E. Neely. 1974. Injury and yield responses of soybeans
to chronic doses of ozone and sulfur dioxide in the field. Phytopathology 64:
132-136.
Heagle, A.S., D.E. Body, and E.K. Punds. 1972. Effect of ozone on yield of sweet corn.
Phytopathology 62: 583-687.
Heck, W.W., R.H. Daines, and I.J. Hindawi. 1970. Other phytotoxic pollutants. In:
Recognition of Air Pollution Injury to Vegetation: a Pictorial Atlas, eds. J.
Jacobson and A.C. Hill. Pittsburgh, Pennsylvania: Air Pollution Control
Association, pp. F1-F24.
Heggestad, H.E. 1973. Photochemical air pollution injury to potatoes in the Atlantic
coastal states. American Potato Journal 50: 315-328.
Heggestad, H.E. and J.H. Bennett. 1981. Photochemical oxidants potentiate yield losses
in snap beans attributable to sulfur dioxide. Science 213: 1008-1010.
Hill, A.C. and J.H. Bennett. 1970. Inhibition of apparent photosynthesis by nitrogen
oxides. Atmospheric Environment 4: 341-348.
Hill, A.C, S. Hill, C. Lamb, and T.W. Barrett. 1974. Sensitivity of native desert
vegetation to SO2 and to SO2 and NO2 combined. Journal of the Air Pollution
Control Association 24: 153-157.
Hill, A.C, H.E. Heggestad, and S.L. Linzon. 1970. Ozone. In: Recognition of Air Pol-
lution Injury to Vegetation: a Pictorial Atlas, eds. J.S. Jacobson and A.C.
Hill. Pittsburgh, Pennsylvania: Air Pollution Control Association, pp. B1-B22.
Hindawi, I.J., J. A. Rea, W.L. Griffis. 1980. Response of bush bean exposed to acid mist.
American Journal of Botany 67(2): 168-172.
Hoffman, G.J., E.V. Mass, and S.L. Rawlins. 1973. Salinity-ozone interactive effects on
yield and water relations of pinto bean. Journal of Environmental Quality 2:
148-152.
Horsman, D.C, A.O. Nicholls, and D.M. Calder. 1980. Growth responses of Dactylis glom-
erata , Lol ium perenne and Phalari s aquatica to chronic ozone exposure.
Australian Journal of Plant Physiology 7: 511-517.
Howell, R.K. and J.H. Graham. 1977. Interaction of ozone and bacterial leafspot of
alfalfa. Plant Disease Reporter 61: 565-567.
Hughes, P.R. 1983. Effects of air pollution on plant -feedi ng insects. Illinois climate:
trends, impacts and issues. In: Proceedings of the 12th Annual ENR Conference,
UrbanaChampaign, Illinois, 1983 September 13-14, Document no. 84/5, Illinois
Department of Energy and Natural Resources, pp. 104-116.
Hughes, P.R. and J. A. Laurence. 1983. Relationship of biochemical effects of air pol-
lutants on plants to environmental problems: insect and microbial interactions.
In: Gaseous Air Pollutants and Plant Metabolism. eds. M.J. Koziol and F.R.
Whatley. London: Butterworth Scientific, pp. 361-377.
Hughes, P.R., A.I. Dickie, and M.A. Penton. 1983. Increased success of the Mexican bean
beetle on field-grown soybeans exposed to sulphur dioxide. Journal of Environ-
mental Quality 12(4): 565-568.
Hughes, P.R., J.E. Potter, and L.H. Weinstein. 1982. Effects of air pollutants on plant-
insect interactions: increases susceptibility of greenhouse-grown soybeans to
the Mexican bean beetle after plant exposure to SO2. Environmental Entomology
11(1): 173-176.
97
Hughes, P.R., J.E. Potter, and L.H. Weinstein. 1981. Effects of air pollutants on plant-
insect interactions: reactions of the Mexican bean beetle to SOz-f umigated
pinto beans. Environmental Entomology 10: 741-744.
International Electric Research Exchange. 1981. Effects of SO2 and its derivatives on
health and ecology. Volume 2 - Natural ecosystems, agriculture and fisheries.
291 pp.
Irving, P.M. 1983. Acidic precipitation effects on crops: a review and analysis of
research. Journal of Environmental Quality 12: 442-453.
Irving, P.M. and J.E. Miller. 1981. Productivity of field-grown soybeans exposed to acid
rain and sulfur dioxide alone and in combination. Journal of Environmental
Quality 10: 473-478.
Irving, P.M., J.E. Miller, and P.B. Xerikos. 1982. The effect of NO2 and SO2. alone and
in combination, on the productivity of field-grown soybeans. In: Air Pollution
by Nitrogen Oxides, eds. T. Schneider and L. Grant. Amsterdam: Elsevier
Scientific Publishing, pp. 521-531.
Jacobson, J.S. 1984. Effects of acidic aerosol, fog, mist and rain on crops and trees.
Philosophical Transactions of the Royal Society of London 305: 327-338.
Jacobson, J.S. 1982. Ozone and the growth and productivity of agricultural crops. In.*
Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterworth Scientific, pp. 293-304.
Jacobson, J.S. and P. Van Leuken. 1977. Effects of acidic precipitation on vegetation.
In: Proceedings of the Fourth International Clean Air Congress. 1977 May
16-20; Tokyo, Japan; pp. 124-127.
Johnston, J.W., D.S. Shriner, C.I. Klarer and D.M. Lodge. 1981. Effect of rain pH on
senescence, growth, and yield of bush bean. Environmental and Experimental
Botany 22(3): 329-337.
Johnston, J.W., R. Dickens, and R.L. Haaland. 1980. Exposure to ozone affects growth of
tall fescue. Highlights Agriculture Research 27:11
Jones, U.S., and E.L. Suarez. 1979. Impact of atmospheric sulfur deposition on agro-
ecosystems In: Atmospheric Sulfur Deposition - Environmental Impact and Health
Effects, eds. D.S. Shriner, C.R. Richmond, and S.E. Lindberg. Ann Arbour: Ann
Arbour Science, pp. 377-396.
Jones, H.C., P.P. Weatherford, and J.C. Noggle. 1979. Power plant siting: assessing
risks of SO2 effects on agriculture. In: 72nd Annual Conference of the Air
Pollution Control Association. 1979 June 24-29; Cincinnati, Ohio.
Katz, M. and G.A. Ledingham. 1939. Effects of SO2 on yield of barley and alfalfa. In:
Effects of SO2 on Vegetation. National Research Council Canada, Publication
No. 815. Ottawa. 457 pp.
Keever, JG.J. and J.S. Jacobson. 1983a. Simulated acid rain effects on zinnia as influ-
enced by available nutrients. Journal of the American Society of Horticultural
Science 108: 80-83.
Keever, JG.J. and J.S. Jacobson. 1983b. Response of Glycine max (L.) Merrill to simu-
lated acid rain: I. Environmental and morphological influences on the foliar
leaching of ^^Rh. Field Crops Research 6: 241-250.
Keever, JG.J. and J.S. Jacobson. 1983c. Response of Glycine max (L.) Merrill to simu-
lated acid rain: II. Localization of foliar injury and growth response. Field
Crops Research 6: 251-259.
Keller, T.H. and J. Muller. 1958. Fortswi ssenschaf tl iche Forschungen 10: 5-63.
Kerr, E.I. and R.A. Reinert. 1968. The response of bean to ozone as related to infec-
tion by Pseudomonas phaseolicola. Phytopathology 58: 1055.
98
Khatamian, H., N.O. Adedipe, and D.P. Ormrod. 1973. Soil-plant-water aspects of ozone
phytotoxicity in tomato plants. Plant and Soil 38: 531-541.
Kochhar, M., U. Blum, and R.A. Reinert. 1980. Effects of Oa and (or) fescue on ladino
clover: interactions. Canadian Journal of Botany 58: 241-249.
Koziol, M.J. and C.F. Jordan. 1978. Changes in carbohydrate levels in red kidney bean
(Phaseolus vulgaris L.) exposed to sulphur dioxide. Journal of Experimental
Botany 29: 1037-1041.
Kratky, B.A., E.T. Fukunaga, J.W. Hylin, and R.T. Kakano. 1974. Volcanic air pollution:
deleterious effects on tomatoes. Journal of Environmental Quality 3: 138-140.
Krause, C.R. and T.C. Weidensaul. 1978a. Effects of ozone on the sporulations , germina-
tion, and pathogenicity of Botrytis cinerea. Phytopathology 68: 195-198.
Krause, C.R. and T.C. Weidensaul. 1978b. Ultrastructural effect of ozone on the host-
parasite relation of Botrytis cinerea and Pelargonium hortorum. Phytopathology
68: 301-307.
Kress, L.W. and J.E. Miller. 1981. Impact of ozone on soybean yield. Argonne National
Laboratory. Radiological and Environal Research Division Annual Report, Part
III Ecological. 1980 January-December, pp. 11-14.
Kress, L.W. and J.M. Skelly. 1982. Response of several eastern forest tree species to
chronic doses of ozone and nitrogen dioxide. Plant Disease Reporter 66:
1149-1152.
Krupa, S.V., and R.N. Kickert. 1987. An Analysis of Numerical Models of Air Pollutant
Exposure and Vegetation Response. Prep for the Acid Deposition Research
Program by the Department of Plant Pathology, University of Minnesota, St.
Paul, Minnesota, U.S.A. and Consultant, Corvallis, Oregon, U.S.A.
ADRP-B-10-87. 113 pp.
Krupa, S.V., J. P. Lodge, Jr., M. Nosal, and G.E. McVehil. 1987. Characteristics of aero-
sols and rain chemistry in north central U.S.A. Proceedings of International
Conference on Acidic Rain: Scientific and Technical Advances, Lisbon,
Portugal, eds. R. Perry, R.M. Harrison, J.N.B. Bell, and J.N. Lester. London:
Selper Ltd, pp. 121-128.
Laurence, J. A. 1981. Effects of air pollutants on plant pathogen interactions. Zeit-
schrift Pflanzenkr Pf lanzenschutz 88(2-3): 156-172.
Laurence, J. A. 1979. Response of maize and wheat to sulfur dioxide. Plant Disease
Reporter 63: 468-471 .
Laurence, J. A., L.H. Weinstein, D.C. McCune, and A.L. Alusio. 1979a. Effects of SO2 on
southern corn leaf blight of maize and stem rust of wheat. Plant Disease
Reporter 63: 975-978.
Laurence, J. A., L.H. Weinstein, A.L. Aluisio, and D.C. McCune. 1979b. Effects of sulfur
dioxide on three plant-virus combinations. Phytopathology 69: 1036.
Laurence, J. A. and F.A. Wood. 1978a. Effects of ozone on infection of soybean by Pseudo-
monas glycinea. Phytopathology 68: 441-445.
Laurence, J. A. and F.A. Wood. 1978b. Effects of ozone on infection of wild strawberry
by Xanthomonas f ragariae. Phytopathology 68: 689-692.
Laurence, J. A., P.R. Hughes, L.H. Weinstein, G.T. Geballe, and W.H. Smith. 1983. Impact
of air pollution on plant-pest interactions: implications of current research
and strategies for future studies. Ecosystems Research Center, Cornell Univer-
sity, Ithaca, NY 14853. ERC Report No. 20. 66 pp.
Lee, J.J. 1981. Acid rain effects on California crop yield and foliar injury. California
Symposium on Acid Precipitation. 1981 January 14-15; San Francisco, California;
pp. 221-232.
99
Lee, J.J. and G.E. Neely. 1980. CERL-OSU acid rain crop study progress report. Air
Pollution Effects Branch, Corvallis Environmental Research Lab., US Environ-
mental Protection Agency, Corvallis, Oregon. 20 pp.
Lee, J.J., G.E. Neely, S.C. Perrigan, and L.C. Grothaus. 1981. Effect of simulated sul-
phuric acid rain on yield, growth and foliar injury of several crops. Environ-
mental and Experimental Botany 21: 171-185.
Lefohn, A.S. and R.W. Brocksen. 1984. Acid rain effects research - a status report.
Journal of the Air Pollution Control Association 34: 1005-1013.
Legge, A.H., E.M. van Zinderen Bakker,Jr., E. Peake, and D.C. Lindsay. 1980. The oxides
of nitrogen and their interactions in the environment: a review. A Final Report
Submitted to the Research Secretariat, Alberta Environment and Canadian Petrol-
eum Association by Kananaskis Centre for Environmental Research, University of
Calgary. Calgary, Alberta. 169 pp.
Linzon, S.N. 1978. Effects of airborne sulfur pollutants on plants. In: Sulfur in the
Environment Part II: Ecological Impacts, ed. J.O. Nriagu. New York: John Wiley
and Sons, pp. 109-162.
Maas, E.V., G.J. Hoffman, and G. Ogata. 1973. Salinity-ozone interactions on pinto bean:
Integrated response to ozone concentration and duration. Journal Environmental
Quality 2: 400-404.
Maly, V. 1974. Scientia Agriculturae Bohemoslovacae, 6(XXIII): 147-159.
Mandl, R.H., L.H. Weinstein, D.C. McCune, and M. Keveny. 1973. A cylindrical, open-top
chamber for the exposure of plants to air pollutants in the field. Journal of
Environmental Quality, 2: 371-376.
Manning, W.J. 1978. Chronic foliar ozone injury: Effects on plant root development and
possible consequences. California Air Environment 7: 3-4
Manning, W.J., W.A. Feder, and I. Perkins. 1972. Effects of Boytrytis and ozone on
bracts and flowers of poinsettia cultivars. Plant Disease Reporter 56: 814-816.
Manning, W.J., W.A. Feder, P.M. Papia, and I. Perkins. 1971a. Influence of foliar ozone
injury on root development and root surface fungi of pinto bean plants.
Environmental Pollution 1: 305-312.
Manning, W.J., W.A, Feder, P.M. Papia, and I. Perkins. 1971b. Effect of low levels of
ozone on growth and susceptibility of cabbage plants to Fusarium oxysporum F.
sp. conglutinans . Plant Disease Reporter 55: 47-49.
Manning, W.J., W.A. Feder, and I. Perkins. 1970a. Ozone injury increases infection of
geranium leaves by Botrytis cinerea. Phytopathology 60: 669-670.
Manning, W.J., W.A. Feder, and I. Perkins. 1970b. Ozone and infection of geranium flowers
by Botrytis cinerea. Phytopathology 60: 1302.
Manning, W.J., W.A. Feder, I. Perkins, and M. Glickman. 1969. Ozone injury and infec-
tion of potato leaves by Botrytis cinerea. Plant Disease Reporter 53: 691-693.
Masaru, N.F., F. Katsuhisa, T. Sankichi, and W. Yutaka. 1980. Effects of inorganic com-
ponents in acid rain on tube elongation of Camellia pollen. Environmental
Pollution A 21 : 51-57.
Masaru, N.F., F. Syozo, and K. Saubro. 1976. Effects of exposure to various injurious
gases on germination of Lily pollen. Environmental Pollution 11: 181-187.
McFee, W.W. 1983. Sensitivity ratings of soils to acid deposition: a review. Envi-
ronmental and Experimental Botany 23(3): 203-210.
McFee, W.W. 1980. Effects of atmospheric pollutants on soils In: Polluted Rain, eds.
T. Y. Toribara, M. W. Miller, and P. E. Morrow. New York: Plenum Press, pp.
307-323.
100
McLaughlin, S.B., R.K. Shriner, S.B. McConathy, and L.K. Mann. 1979. The effects of SO2
dosage kinetics and exposure frequency on photosynthesis and transpiration of
kidney beans ( Phaseolus vulgaris L.). Environmental and Experimental Botany
19: 179-191.
Morrison, I.K. 1984. Acid rain: a review of literature on acid deposition effects on
forest ecosystems. Forestry Abstracts 45(8): 483-506.
Mortvedt, J.J. 1983. Impacts of acid deposition on micronutrient cycling in agro-
ecosystems. Environmental and Experimental Botany 23(3): 24.
Moyer, J.W. and S.H. Smith. 1975. Oxidant injury reduction on tobacco induced by tobacco
etch virus infection. Environmental Pollution 9: 103-106.
Mudd, J.B. 1973. Biochemical effects of some air pollutants on plants. In: Air pol-
lution damage to vegetation, ed. J. A. Naegele. Advances in Chemistry Series,
122: 31-47.
Mudd, J.B. and T.T. Kozlowski. 1975. Responses of Plants to Air Pollution. New York:
Academic Press. 383 pp.
National Academy of Sciences. U.S. 1977a. Ozone and other photochemical oxidants.
Committee on Medical and Biological Effects of Environmental Pollutants.
Washington, D.C. 719 pp.
National Academy of Sciences. U.S. 1977b. Nitrogen oxides. Committee on Medical and
Biological Effects of Environmental Pollutants. Washington, D.C. 333 pp.
Neely, G.E., D.T. Tingey, and R.G. Wilhour. 1977. Effects of ozone and sulfur dioxide
singly and in combination on yield, quality, and N-fixation of alfalfa. In:
International Conference on Photochemical Oxidant Pollution and its Control,
ed. B. Dimitriades. EPA-600/3-77-001 b, US Environmental Protection Agency,
Research Triangle Park, North Carolina, pp. 663-673.
Neufeld, H.S., J. A. Jernsted, and B.L. Haines. 1985a. Direct foliar effects of simulated
acid rain. I. Damage, growth and gas exchange. New Phytologist 99: 389-405.
Neufeld, H.S., J. A. Jernsted, and B.L. Haines. 1985b. Direct foliar effects of simulated
acid rain. II. Leaf surface characteristics. New Phytologist 99: 407-416.
Neumann, P.M., Y. Ehrenreich, and Z. Golab. 1981. Foliar fertilized damage to corn
leaves: relation to cuticular penetration. Agronomy Journal 73: 979-982.
Noggle, J.C. and H.C. Jones. 1982. Effects of air pollutants on wheat and grains. In:
Effects of Air Pollution on Farm Commodities, eds. J.S. Jacobson and A. A.
Millen. Arlington: The Izaak Walton League, pp. 79-90.
Norby, R.J. and R.J. Luxmoore. 1983. Growth analysis of soybean exposed to simulated
acid rain and gaseous air pollutants.* New Phytologist 95(2): 277-287.
Oertli, J.J., J. Harr, and R. Guggenheim. 1977. The pH values as an indicator for the
leaf surface microenvironment. Zeitschift fur Pf lanzenkrankhei ten und Phlanzen-
schutz 84: 729-737.
Ogata, G. and E.V. Maas. 1973. Interactive effects of salinity and ozone on growth
and yield of garden beet. Journal of Environmental Quality 2: 518-520.
Ormrod, D.P. 1984. Assessing effects of pollutant mixtures (Oa, SO2, NO2): methods and
knowledge - crops. In: Proceedings of the 77th Annual Meeting of the Air Pol-
lution Control Association, 1984 June 24-29; San Francisco, California. 11 pp.
Ormrod, D.P., N.O. Adedipe, and G. Hofstra. 1971. Responses of cucumber, onion, and
potato cultivars to ozone. Canadian Journal of Plant Science 51: 283-288.
Oshima, R.J. 1973. Effect of ozone on a commercial sweet corn variety. Plant Disease
Reporter 57: 719-723.
101
Paparozzi, E.T. 1981. The effects of simulated acid precipitation on leaves of Betula
al leqhniensis Britt. and Phaseolus vulgaris cv. red kidney. Ithaca, New York:
Cornell University. 127 pp. PhD. Thesis.
Pell, E.J., W.C. Weissberger, and J.J. Speroni. 1980. Impact of ozone on quantity and
quality of greenhouse -grown potato plants. Environmental Science and Technology
14: 568-571.
Pell, E.J., F.L. Lukfzic, R.G. Levine, and W.C. Weisberger. 1977. Response of soybean
foliage to reciprocal challenges by ozone and a hypersensitive-response-inducing
Pseudomonad. Phytopathology 67: 1342-1345.
Perchorowicz, J.R. and I. P. Ting. 1974. Ozone effect on plant cell permeability. Ameri-
can Journal of Botany 61(7): 787-793.
Plocher, M.D., S.C. Perrigan, R.J. Hevel, R.M. Cooper, and D.N. Moss. 1985. Simulated
sulphuric acid rain on crops. Agricultural Experiment Station, Oregon State
University, Corvallis, Oregon Special Report 739. 64 pp.
Pratt, B.C., M.R. Coscio, D.W. Gardner, B.I. Chevone, and S.V. Krupa. 1983. An analysis
of the chemical properties of rain in Minnesota. Atmospheric Environment 17:
347-355.
Prince, R. and F.F. Ross. 1972. Sulphur in air and soil. Water, Air, and Soil Pollu-
tion 1 : 286-302.
Proctor, J.T.A. 1983. Effect of simulated sulphuric acid rain on apple tree foliage,
nutrient content, yield and fruit quality. Environmental and Experimental
Botany 23(2): 167-174.
Reinert, R.A. 1984. Plant response to air pollutant mixtures. Annual Review of Phyto-
pathology 22: 421-442.
Reinert, R.A. and G.V. Gooding, Jr. 1978. Effect of ozone and tobacco streak virus alone
and in combination on Nicotiana tabacum. Phytopathology 68: 15-17.
Reinert, R.A. and J.S. Sanders. 1982. Growth of radish and marigold following repeated
exposure to nitrogen dioxide, sulfur dioxide, and ozone. Plant Disease 66:
122-124.
Runeckles, V.C., K. Palmer, and K. Giles. 1978. Effects of sequential exposures to NO2
and 03 on plants. In: Proceedings of the Third International Congress of
Plant Pathology, Munich, Abstracts; p. 343.
Russell, E.W. 1973. Soil Conditions and Plant Growth. New York: Longman. 849 pp.
Sandhu, H.S., H.P. Sims, R.A. Hursey, W.R. MacDonald, and B.R. Hammond. 1980. Environ-
mental Sulphur Research in Alberta: a Review. Research Secretariat. Edmonton:
Alberta Department of Environment. 90 pp.
Schultz, J.C. and I.T. Baldwin. 1982. Oak leaf quality declines in response to defolia-
tion by gypsy moth larvae. Science 217(4555): 149-151.
Shannon, J.G. and C.L. Mulchi. 1974. Ozone damage to wheat varieties at anthesis. Crop
Science 14:335-337.
Shertz, R.D., W.J. Kender, and R.C. Musselman. 1980. Effects of ozone and sulfur dioxide
on grapevines. Science and Horticulture 13: 37-45.
Shew, B.B., R.A. Reinert, and K.R. Barker. 1982. Response of tomatoes to ozone, sulfur
dioxide, and infection by Pratylenchus penetrans . Phytopathology 72: 822-826.
Shinohara, T., Y. Yamamoto, H. Kitano, and M. Fukuda. 1974. Interactions of light and
ozone injury in tobacco. In: Proceedings of Crop Science Society of Japan 43:
433-438.
Shriner, D.S. 1983. Interaction of gaseous pollutant and acid rain effects. Acid Rain
Program Review Meeting. February 21; Raleigh, North Carolina; 91 pp.
102
Shriner, D.S. 1980. Vegetation surfaces: a platform for pollutant/parasite interactions.
In: Proceedings of the 12th University of Rochester Symposium on Environmental
Toxicity: Polluted Rain, eds. J.M. Wood, T.Y. Toribara, R.R. Miller, and P.E.
Morrow. 1979 May 21-23; Rochester, New York. Environmental Science Research
Series, New York: Plenum Press; pp. 259-270.
Shriner, D.S. 1978. Interactions between acidic precipitation and SO2 or Oa: effects
on plant response. Phytopathology News 12: 153.
Shriner, D.S. 1977. Effects of simulated rain acidified with sulfuric acid on host-
parasite interactions. Water, Air, and Soil Pollution 8: 9-14.
Shriner, D.S. 1974. Effects of simulated rain acidified with sulfuric acid on host-
parasite interactions. Raleigh, N. Carolina: North Carolina State University.
79 pp. Ph.D. Thesis.
Shriner, D.S. and J.W. Johnston. 1981. Effects of simulated, acidified rain on nodula-
tion of leguminous plants by Rhizobium spp. Environmental and Experimental
Botany 21 : 199-209.
Shriner, D.S. and E.B. Cowling. 1980. Effects of rainfall acidification on plant patho-
gens. In: Effects of Acid Precipitation on Terrestrial Ecosystems, eds. T.C.
Hutchinson and M. Havas. New York: Plenum Press, pp. 435-442.
Sparrow, A.H. and L.A. Schairer. 1974. Mutagenic response of Tradescantia to treatment
with X-rays, IMS, DBE, ozone, SO2, N2 and several insecticides. Mutation
Research 26: 445.
Spierings, F.H.F.G. 1971. Influence of fumigations with nitrogen dioxide on growth and
yield of tomato plants. Netherland Journal of Plant Pathology 77: 194-200.
Stanford Research Institute. 1981. An estimate of the non-health benefits of meeting the
Secondary National Ambient Air Quality Standards. Final Report. National
Commission on Air Quality, Washington, D.C. SRI Project 2094; Menlo Park,
California, 53 pp.
Strong, W.L. and K.R. Leggat. 1981. Ecoregions of Alberta. Alberta Energy and Natural
Resources. Report Number T/4. 64 pp.
Tamm, CO. and E.B. Cowling. 1976. Acidic precipitation and forest vegetation. General
Technical Report, Northeastern Forest Experiment Station, USDA Forest Service.
No. NE-23. pp. 1011-1024.
Taniyama, T. 1972. Bulletin of the Faculty of Agriculture, Mie University, Tsu, Japan
44: 11-130.
Taniyama, T., H. Arikado, and Y. Iwata. 1972. Studies on the mechanism of injurious
effects of toxic gases on crop plants. Proceedings of the Crop Science Society
of Japan 40(4): 455-461.
Taylor, O.C. 1984. Organismal responses of higher plants to atmospheric pollutants:
photochemical and other. In: Air Pollution and Plant Life, ed. M. Treshow.
Chichester: John Wiley and Sons. pp. 215-238.
Taylor, O.C. and D.C. Maclean. 1970. Nitrogen oxides and the peroxyacyl nitrates. In:
Recognition of Air Pollution Injury to Vegetation: A Pictorial Atlas, eds.
J.S. Jacobson and A.C. Hill. Pittsburg, Pennsylvania: Air Pollution Control
Administration, pp. E1-E14.
Taylor, O.C, CR. Thompson, D.T. Tingey, and R.A. Reinert. 1975. Oxides of nitrogen.
In: Responses of Plants to Air Pollution, eds. J.B. Mudd and T.T. Kozlowski.
New York: Academic Press, pp. 121-139.
Thomas, M.D. 1935. Absorption of sulfur dioxide by alfalfa and its relation to leaf
injury. Plant Physiology 10: 291-307.
Thompson, CR. and G. Kats. 1978. Effects of continuous H2S fumigation on crop and for-
est plants. Environmental Science and Technology 12: 550.
103
Thompson, C.R., G. Kats, and R.W. Lennox. 1980. Effects of sulfur dioxide and/or nitrogen
dioxide on native plants of the Mojave Desert and Eastern Mojave-Colorado
Desert. Journal of the Air Pollution Control Association 30: 1304-1309.
Tingey, D.T. 1977. Ozone induced alterations in plant growth and metabolism. In: Pro-
ceedings of the Second International Conference on Photochemical Oxidant
Pollution and Its Control, ed. B. Dimitriades. Research Triangle Park, North
Carolina: Environmental Protection Agency, pp. 601-609.
Tingey, D.T. and U. Blum. 1973. Effects of ozone on soybean nodules. Journal of Envi-
ronmental Quality 2: 341-342.
Tingey, D.T., J. A. Dunning, and G.M. Jividen. 1973a. Radish root growth reduced by acute
ozone exposures. In: Proceedings of the Third International Clean Air Con-
gress. 1973 October 8-12; Dusseldorf, Germany; pp. A154-A156.
Tingey, D.T., R.C. Fites, and C. Wickliff. 1973b. Foliar sensitivity of soybeans to ozone
as related to several leaf parameters. Environmental Pollution 4: 183-192.
Tingey, O.T., W.W. Heck, C. Wickliff, and R.A. Reinert. 1973c. Chronic ozone or sulfur
dioxide exposures, or both, affect the early vegetative growth of soybean.
Canadian Journal of Plant Science 53: 875-879.
Tingey, D.T., W.W. Heck, and R.A. Reinert. 1971a. Effect of low concentrations of ozone
and sulfur dioxide on foliage, growth and yield of radish. Journal of the
American Society of Horticultural Science 96: 369-371.
Tingey, D.T., W.W. Heck, J. A. Dunning, and R.A. Reinert. 1971b. Vegetation injury from
the interaction of nitrogen dioxide and sulfur dioxide. Phytopathology 61:
1506-1511 .
Torn, M.S., J.E. Degrange, and J.H. Shinn. 1987. The Effects of Acidic Deposition on
Alberta Agriculture: A Review. Prep for the Acidic Deposition Research
Program by the Environmental Sciences Division, Lawrence Livermore National
Laboratory. ADRP-B-08/87 . 175 pp.
Treshow, M., 6. Dean, and F.M. Harner. 1967. Stimulation of tobacco mosaic virus-induced
lesions on bean by fluoride. Phytopathology 57: 756-758.
Troiano, J., J.S. Jacobson, and L. Heller. 1984. Effects of simulated acidic rain applied
alone and in combination with ambient rain on growth and yield of field-grown
snap bean. Agriculture, Ecosystems and Environment 11: 161-172.
Troiano, J., L. Colavito, L. Heller, D.C. McCune and J.S. Jacobson. 1983. Effects of
acidity of simulated rain and its joint action with ambient ozone on measures
of biomass and yield in soybean. Environmental and Experimental Botany 23:
113-119.
Troiano, J., L. Heller, and J.S. Jacobson. 1982. Effect of added water and acidity of
simulated rain on growth of field-grown radish. Environmental Pollution,
Series A 29(1): 1-11 .
United States National Research Council. 1983. Acid Deposition: Atmospheric Processes
in Eastern North America. National Academy Press, Washington, D.C. 375 pp.
US Environmental Protection Agency. 1978. Effects of photochemical oxidants on vegeta-
tion and certain microorganisms. In: Air Quality Criteria for Ozone and Other
Photochemical Oxidants, pp. 253-259.
Van Haut, H. and H. Stratmann. 1967. Experimentel le Untersuchungen uber die Wirkung von
Stickstof fdioxid auf Pflanzen. Immissions Bodennutzungssch des Landes Nordrhein-
Westfalen 7: 50-70.
Warteresiewicz, M. 1979. Archiv Ochrony Srodowiska 7: 95-166. (Original not seen; taken
from Godzik and Krupa 1982).
104
Wedding, R.T. and L.C. Erickson. 1955. Changes in the permeability of plant cells to
and water as a result of exposure to ozonated hexene (smog). American
Journal of Botany 42: 570-575.
Whitmore, M.E. and P.H. Freer-Smith. 1982. Growth effects of SO2 and/or NO2 on woody
plants and grasses during spring and summer. Nature 300: 55-57.
Whitmore, M.E. and T.A. Mansfield. 1983. Effects of long-term exposures to SO2 and NO2
on Poa pratensis and other grasses. Environmental Pollution 31: 217-235.
Zahn, R. 1961. Wirkung von Schwefel dioxid auf die vegetation, Ergebnisse aus Begasungs-
versuchen. Staub 21: 56-60.
Zeevaart, A.J. 1976. Some effects of fumigating plants for short periods of time with
NO2. Environmental Pollution 11: 97-108.
Ziegler, I. 1975. The effect of SO2 pollution on plant metabolism. Residue Review 56:
79-105.
105
4. NUMERICAL MODELS OF AIR POLLUTANT EXPOSURE AND VEGETATION RESPONSE
The following section deals with various types of models that attempt to define
and predict the effects of air pollutant exposure on plants. Most of these models use
dose-response relationships to define and predict effects. Descriptions of the various
models may be found in Krupa and Kickert (1987).
4.1 TYPES OF MODELS
A model can be described as information, data, principles and the like, arranged
or grouped, usually mathematically, so as to represent or describe a certain idea or
condition. Models of biological systems are generally grouped into one of two types
(Woodmansee 1974) :
a. Statistical or empirical models are composed of mathematical equations
that have been statistically derived from sets of data collected in the
field or laboratory. These types of models describe the statistical
relationships between the dependent variable and one or more independent
variables.
b. Mechanistic or process models attempt to represent a biological system in
terms of basic, well defined laws or relationships.
Both types of models have advantages and limitations. Statistical models are
relatively simple, have easily accessible input requirements and have precision in their
output but they have low realism, scientific value, and general applicability to loca-
tions other than those in which they were developed. Conversely, mechanistic models are
more realistic and are of scientific value but they are usually complex, have difficult
to acquire input requirements, and have low precision in their output.
In evaluating the impacts of air pollutants on cultivated and natural vegeta-
tion, a number of investigators have proposed statistical models (Larsen and Heck 1976;
Benson et al. 1982; Heck et al. 1982; Loehman and Wilkinson 1983; Medeiros et al. 1983;
Nosal 1983, 1984; Fox et al . 1986). Others have developed mechanistic models (Ares
1979; Andersson et al . 1980; Haines and Waide 1980; Kercher 1980; Luxmoore 1980; West
etal. 1980; Coughenour 1981; Heasley etal. 1981; Kercher and Axelrod 1981; Miller
et al. 1982; Harwell and Weinstein 1983; King et al. 1983; Mortensen 1984). These and
other similar models have attempted to explain the relationships between acute or
chronic pollutant exposures and vegetation responses.
4.2 ACUTE VERSUS CHRONIC EXPOSURE
In the present context, "acute exposure" is defined as the occurrence of
short-term (hours to days) high pollutant concentrations. Chronic exposure is defined
as the occurrence of long-term (weeks, months, entire life cycle) low pollutant concen-
trations with periodic intermittent "episodes" or peaks. Generally, acute exposures
cause symptoms of injury from which the plant may or may not recover depending on the
timing and magnitude of the stress. On the other hand, chronic exposures may or may not
cause visible symptoms of injury; however, chronic effects can result in changes in the
growth, reproduction, productivity, and quality of the plants.
106
4.3 CHARACTERISTICS OF AMBIENT AIR QUALITY
When creating a numerical model to describe vegetation responses to air pollut-
ant exposure "understanding of the nature of the cause is as important as understanding
the nature of the receptor response" and "sound integration of the two aspects is
critical" (Krupa and Kickert 1987). Unfortunately, authors of many models have over-
simplified the nature of ambient air quality and therefore the results of their models
are of questionable value. Development of models with good predictive capabilities
requires that the authors correctly describe the ambient (real world) air quality.
In evaluating the effects of pollutants under real world conditions, one must
define the spatial and temporal variability in the pollutant concentrations. For
example, high concentrations of ozone (Oa) are generally observed during the daylight
hours, while high concentrations of fine particulate sulphate (S04^~) are observed
during night-time hours (Stevens et al. 1978). This temporal variability in the patterns
of the two pollutants has been shown to be of importance in vegetation effects. Fine
particulate sulphate (S04^ ) alone did not produce visible injury, but when plants
were exposed to fine particulate sulphate (S04^ ) followed by ozone (O3) more than
additive visible Oa-type injury was observed (Herzfeld 1982; Chevone et al. 1986).
In addressing this issue one must also correctly provide an appropriate numeri-
cal description of the frequency distribution of the pollutant concentrations and
exposure over time. Many scientists assume that, for the purposes of their models, the
frequency distribution of pollutant concentration follows a normal, "bel 1 -shaped"
distribution (e.g., Oshima et al. 1976; Male 1982; and Heagle et al. 1986). In reality
it has been shown that the frequency distribution of the occurrence of oxides of nitrogen
(Pratt et al. 1983), sulphur dioxide (Berger et al. 1982; Fowler and Cape 1982; and
Buttazzoni et al. 1986) and ozone (Lefohn and Benedict 1982; Nosal 1984) are all skewed
toward low concentrations, with a long tail toward high values. Several mathematically
derived theoretical distributions have been proposed to describe the nature of the
observed pollutant frequency distributions, for example, the log-normal distribution
(Fowler and Cape 1982; Male 1982) and the two parameter gamma distribution (Berger
et al. 1982). However, most recently the Weibull distribution has been proposed as a
universal form for the interpretation of air quality data (Krupa and Kickert, 1987)
because the log-normal distribution does not possess the characteristics to sufficiently
explain the generality or universality and the gamma distribution is computationally
inconvenient.
In addition to the previously discussed gaseous pollutants (ozone and oxides of
nitrogen and sulphur) the phenomenon of acidic rain is of substantial concern. Almost
all studies of acidic rain relate to the use of "simulated rain." This involves treating
plants with a "solution of constant chemical composition, applied artificially at
constant or varying amounts." However, in the real world the chemistry of precipitation
varies significantly within and between individual rain events (Pratt et al. 1983).
This fact questions the results of the "simulated rain" experiments and the models
derived from the results of such experiments. The experiments normally use "average"
values of the constituents in precipitation to develop the "simulated rain." The use of
average values and linear statistical regression techniques are inappropriate because
the data invariably violate the statistical assumptions of normality, homoscedasticity,
unbiased residuals, and independence (Nosal and Krupa 1986). In addition, there are a
107
number of other concerns introduced through artifacts caused by sampling methodology
(Electric Power Research Institute 1986), the duration of the sample collection period
(Sisterson et al. 1985), the methods used in data analysis (Krupa and Kickert 1987), and
the seasonality of data collection.
Finally, Krupa and Kickert (1987) raised concerns that little if any effort has
been made to adequately describe the statistical distributions of the fine particles
which contain a major portion of sulphate and nitrate.
4.4 THE CONCEPT OF POLLUTANT DOSE
Pollutant dose can be defined as the exact amount of a given pollutant to which
a given receptor (plant) is subjected at one time, or at staged intervals. Historically,
dose has been expressed as:
1. Pollutant concentration multiplied by the duration of the exposure;
2. Pollutant concentration divided by the duration of the exposure; or,
3. The sum of the concentration multiplied by the duration of the exposure,
when the pollutant concentration exceeds a set minimum threshold (inte-
grated exposure: Lefohn and Benedict 1982).
The pollutant dose "indices" have several drawbacks. They assume a normal
distribution of ambient pollutant concentrations. The negative consequences of this
assumption are illustrated by the work of Hogsett et al. (1985). Further problems with
pollutant averaging techniques have been pointed out by Krupa and Gardner (in prepara-
tion). They found that hourly averages of sulphur dioxide concentration, downwind of a
large point source in Minnesota, were reported to be zero during 90% of the ten year
monitoring period. However, when the continuous data were examined, within the 90% of
the time numerous instances of sulphur dioxide concentrations of 0.50 ppm or higher were
observed.
The matter is further confounded by the fact that exposure to low concentrations
of pollutant can either predispose plants to injury from subsequent episodes oi* can
confer tolerance (Godzik and Krupa 1982). In addition, the developmental stage of the
plant during the time of exposure can have a marked effect on its response to a pollutant
dose (Lockwood et al. 1977; Teng and Gaunt 1980; and Benson et al. 1982). Finally,
atmospheric factors (light, temperature, relative humidity, and carbon dioxide), edaphic
factors (moisture and nutrient availability) and other biological factors (specific
genetics of the plant cultivar and the presence of pests and pathogens) influence the
growth and development of the plant and further confound attempts to determine the
response of the plant to the pollutant exposure.
In the context of air pollutant exposure and plant response a satisfactory
numerical expression of dose should at least consider:
1. the artifacts of pollutant averaging techniques;
2. the episodicity of pollutant occurrence and exposure;
3. the time intervals between episodes; and,
4. the relationship between the pollutant stress and the growth stage of the
plant.
108
Nosal (1983) provided a mixed, multivariate, polynomial, Fourier regression model which
accounts for the number of pollutant episodes, the highest peak in pollutant concentra-
tion, and the cumulative numerical integral (concentration over time) of exposure.
However, the model does not provide a complete numerical explanation of the importance
of the individual pollutant episodes or the relationship of exposure to the growth stage
of the plant.
A final aspect of dose that has not been addressed by any models relates to
"exposure dose" versus "effective dose" (Runeckles 1974). Exposure dose is the pollutant
regime to which the plant is exposed. Effective dose is the actual quantity of the
pollutant which is absorbed by the plant. No field techniques are currently available
to measure effective dose, although the use of radioactive tracers or the computational
approach of pollutant absorbed dose (PAD) suggested by Fowler and Cape (1982) offer
possibilities. [PAD = mean pollutant concentration multiplied by exposure time multiplied
by canopy conductance.]
4.5 MATHEMATICAL MODELS FOR CHARACTERIZING PLANT RESPONSE TO AIR POLLUTANT STRESS
Krupa and Kickert (1987) provide a tabular summary of their review of 30
mathematical models for characterizing plant response to air pollutant stress. This
tabular summary provides basic information on the pollutants and receptors considered by
the models, a brief statement of the major biological paradigm involved in each case,
and the results obtained through the application of each model. In addition, a statement
of the advantages, limitations, applicability, and possible modifications for each model
is provided. Table 27 provides a summary of the models reviewed by Krupa and Kickert
(1987).
4.5.1 Acute Pollutant Exposure and Plant Response Models
Of the 30 models summarized in the review only 6 describe acute pollutant
exposure and plant response. Three of these are statistical (empirical) models and
three are mechanistic (process oriented) models. The three statistical models are
designed to describe the effects of pollutants on agricultural crops. These models are
suitable only for preliminary assessment of plant response or preliminary guideline
preparation. Of the three process oriented models of acute effects, one is suitable for
research only, one requires further development, and one has inadequate documentation.
4.5.2 Chronic Pollutant Exposure and Plant Response Models
The remaining 24 models reviewed by Krupa and Kickert deal with the effects of
chronic exposure of pollutants. Most of the statistical models are also only suitable
for preliminary assessment of yield loss or preliminary guideline preparation (Table
28). The models of Nosal (1983, 1984) are suitable for environmental assessments if a
suitable data base is available. The mechanistic models of response to chronic exposure
are primarily suitable for research only. Many of them suffer from inadequate documen-
tation. The models of King (1986) and Mortensen (1984) could be used as preliminary
guidelines to develop further research programs for the chronic effects of pollutants on
agricultural crops.
109
I— a> ro
^ +-> +->
(O to c
O 3 (U
•r- or e
I— <U =J
-H CL-O O
O Q. C O
z <: t-H Q
I- o
to <U T-
C C -H
•r— T- (O
e »- s-
•r- <U fO
r— -O Q.
O) Q)
i- a i-
CL C3 0_
J- -H
ra c
C Q) T3
•r; E »—
E «/) a>
•r- »/) 'f-
I— <U >- t/»
(U (/I to
I- •/) U- O
Q- <C O -J
(O a>
C +J ro ^
a> c= O)
E <U «0 I—
c E +-> ^
O (/) ra (O
I- I/) T3 I—
•p- Q) •!-
> i/» 4*_ ra
C «>1 'f- >
UJ <C ^ — ^ ro
-O Q.
o >>
ro
v£> C
+J OD ro
oj cy> E
f— -H
O (U
u. 3
CO
I/)
C Jk^ ro C71
(U u c
I/) <U ^ 'r-
J- 3: o »—
ro 0)
(U <U
> N
O) ro
IE
CM
C OS
O <T»
I/) I —
CO
CO
♦ • rO
I — r—
ro . ^ . -O C ro 4->
r- r- CO O)
+-» ro •> ro ro i/> +-»
Q) ro c <l) t/>
00 <uoo 0)00 <Doo «o^ ror- t-oo
I — CTi CTi CTi E I — E Cn CTi
r— CT)r— ^r— Jiir— ^'r-'i-f— CDr—
- — rO^O> — U ^ Qi 'S
O) O) <U O to (U
3: a: 3: — I o s:
o to
•I- O.
i- O
CD l-
<: <-)
o to
•I- Q.
i- o
< o
110
00
r~
<D
"O
o
E
-H
c
O
in
t/i
<v
(_>
o
^
o.
•r-
o
(U
E
a>
■M
o
■H
•r-
r—
•r-
plica
987).
o.
<a
■-—^
u-
o
■M
l-
0)
■M
o
c
•r-
(U
E
to
to
C
Qi
to
rt3
to
CO
ro
O.
0)
JZ
+->
>)
X)
O
TD
>> 0)
3
ro
(U
•1—
>
13
00
CVJ
I
<— c
^ o
o a> +->
•r- +J (O
r- (O +J
O. =J C
O. CT <U
< O) E
T3 13
+-> to O
o c o
2 6-=) Q
C
O) I- Q.
I- <U O
•f- ^ f—
cr s- >
0> 3 <U
ac: u. o
u
lO 0)
•r- Q)
r— -O
Q. CD
+->
1— to
1— C
O) T-
+-> C -1-
>> r- r— to CO
a> CD r— C CD
I— CT> •!- <y>
to ( — 3 a> f—
(0 >— ' i- 3 ^
O) (0
^ 3Z
O
x: ^
C r—
<U (D
C31 f—
O
r—
a>
o
na
T3
CD
r—
fO
r~
fO
■H
3
a>
«o
4-»
et
a%
c
■a
£=
<0
0)
a>
cn
o
03
s-
CO
CVJ
o
1 —
to O
o
■H
O
o
0)
OD
i-
CD
0)
CD
to
CD
to
CD
0)
OD
o
a»
a>
cr»
J-
O)
E
CP 1—
r—
o
1/1
Q)
c
r—
+->
X
c
r—
t-
(U
■o
•r-
to
=J
•r-
•r-
a>
s_
c:
fO
(U
_J
s:
:^
<c
<:
3:
3
vO
CD C
€7) a>
r— to
w C
00 +J
C S-
•r- O
•H
to
+->
c
+->
c:
to
ro
ro
u
ro
ro
+->
i-
r—
<u
l-
^—
o
=3
Ol
M-
3
Q_
<u
+-»
M-
■M
ex.
UJ
>>
3
ro
\-
UJ
o to
1-
U
<->
to
i-
•1- o.
3
•p-
•p-
CL
3
1 —
(U
t- o
C
l-
O
+->
O)
■H
cn J-
ro
o
cn
ro
T3
=J
< CJ
Z
<:
Z
O
o
sz
z:
<c
C_3
Ill
4.6 NUMERICAL MODELS OF POLLUTANT EXPOSURE AND VEGETATION RESPONSE: LITERATURE
CITED.
Andersson, F., T. Fagerstrom, and S.I. Nilsson. 1980. Forest ecosystem responses to acid
deposition - hydrogen ion budget and nitrogen/tree growth model approaches.
In: Effect of Acid Precipitation on Terrestrial Ecosystems, eds. T.C.
Hutchinson and M. Havas. New York: Plenum Press, pp. 319-334.
Ares, J. 1979. Modelling the fate of atmospheric fluoride in a coastal semi-arid region.
I. Systems analysis and identification. In: State-of-the-Art in Ecological
Modelling. Proceedings of the Conference on Ecological Modelling,
Copenhagen, Denmark, ed. S.E. Jorgensen. New York: Pergamon Press, pp.
375-403.
Benson, F.J., S.V. Krupa, P.S. Teng, and D.E. Welsch. 1982. Economic assessment of air
pollution damage to agricultural and si 1 vicultural crops in Minnesota. Final
Report to Minnesota Pollution Control Agency, Roseville, Minnesota. 270 pp.
Berger, A., J.L. Melice, and CI. Demuth. 1982. Statistical distributions of daily and
high atmospheric SO2 concentrations. Atmospheric Environment 16: 2863-2877.
Buttazzoni, C, I. Lavagnini, A. Marani, F.Z. Grandi, and A. Del Turco. 1986. Probability
model for atmospheric sulphur dioxide concentrations in the area of Venice.
Journal of the Air Pollution Control Association 36: 1028-1030.
Chevone, B.I., D.E. Herzfeld, S.V. Krupa, and A.H. Chappelka. 1986. Direct effects of
atmospheric sulfate deposition on vegetation. Journal of the Air Pollution
Control Association 36: 813-816.
Coughenour, M.B. 1981. Relationship of SO2 dry deposition to a grassland sulfur
cycle. Ecological Modelling 13: 1-16.
EPRI, Electric Power Research Institute. 1986. Proceedings: Methods for Acidic Deposi-
tion Measurement. EPRI EA-4663. Electric Power Research Institute, Palo
Alto, California.
Fowler, D. and J.N. Cape. 1982. Air pollutants in agriculture and horticulture. In:
Effects of Gaseous Air Pollution in Agriculture and Horticulture, eds. M.H.
Unsworth and D.P. Ormrod. London: Butterworth Scientific, pp. 3-26.
Fox, C.A., W.B. Kincaid, T.H. Nash, III, D.L. Young, and H.C. Fritts. 1986. Tree ring
variation in Western larch ( Larix occidental i s) exposed to sulfur dioxide
emissions. Canadian Journal of Forest Research 16: 283-292.
Godzik, S. and S.V. Krupa. 1982. Effects of sulfur dioxide on growth and productivity of
crop plants. In: Effects of Gaseous Air Pollution in Agriculture and Horti-
culture, eds. M.H. Unsworth and D.P. Ormrod. London: Butterworth Scientific,
pp. 247-265.
Haines, B. and J. Waide. 1980. Predicting potential impacts of acid rain on elemental
cycling in a Southern Appalachian deciduous forest at Coweeta. Iin: Effect
of Acid Precipitation on Terrestrial Ecosystems, eds. T.C. Hutchinson and M.
Havas. New York: Plenum Press, pp. 335-340.
Harwell, M.A. and D.A. Weinstein. 1983. Modelling the effects of air pollutants on for-
ested ecosystems. In: Analysis of Ecological Systems: State-of-the-Art in
Ecological Modelling, eds. W.K. Lauenroth, G.V. Skogerboe and M. Flug.
Amsterdam: Elsevier Scientific Publishing, pp. 497-502.
Heagle, A.S., V.M. Lesser, J.O. Rawlings, W.W. Heck, and R.B. Philbeck. 1986. Responses
of soybeans to chronic doses of ozone applied as constant or proportional
additions to ambient air. Phytopathology 76: 51-56.
Heasley, J.E., W.K. Lauenroth, and J.L. Dodd. 1981. Systems analysis of potential air
pollution impacts on grassland ecosystems. In: Energy and Ecological Model-
ling, eds. W.J. Mitsch, R.W. Bosserman and J.M. Klopatek. New York:
Elsevier/North Holland, pp. 347-359.
112
Heck, W.W. and D.T. lingey. 1 979. Nitrogen dioxide: time -concentration model to predict
acute foliar injury. US Environmental Protection Agency, EPA-600/3-79-057 .
Corvallis, Oregon. 16 pp.
Heck, W.W., O.C. Taylor, R. Adams, G. Bingham, J. Miller, E. Preston, and L. Weinstein.
1982. Assessment of crop losses from ozone. Journal of the Air Pollution
Control Association 32: 353-361.
Heck, W.W., W.W. Cure, J.O. Rawlings, L.J. Zaragosa, A.S. Heagle, H.E. Heggestad, R.J.
Kohut, L.W. Kress, and P.J. Temple. 1984a. Assessing impacts of ozone on
agricultural crops. I. Overview. Journal of the Air Pollution Control
Association 34: 729-735.
Heck, W.W., W.W. Cure, J.O. Rawlings, L.J. Zaragosa, A.S. Heagle, H.E. Heggestad, R.J.
Kohut, L.W. Kress, and P.J. Temple. 1984b. Assessing impacts of ozone on
agricultural crops. II. Crop yield functions and alternative exposure statis-
tics. Journal of the Air Pollution Control Association 34: 810-817.
Herzfeld, D.E. 1982. Interactive effects of sub-micron sulfuric acid aerosols and ozone
on soybean and pinto bean. St. Paul, Minnesota: University of Minnesota.
105 pp. M.Sc . Thesis.
Hogsett, W.E., D.l. Tingey, and S.R. Holman. 1985. A programmable exposure control sys-
tem for determination of the effects of exposure regimes on plant growth.
Atmospheric Environment 19: 1135-1145.
Kercher, J.R. 1980. Developing realistic crop loss models for air pollution stress. Xn:
Crop Loss Assessment, eds. P.S. Teng and S.V. Krupa. St. Paul, Minnesota:
Agricultural Experimental Station Miscellaneous Publication 7, University of
Minnesota, pp. 90-97.
Kercher, J.R. and M.C. Axelrod. 1981. SILVA: a model for forecasting the effects of
SO2 pollution on growth and succession in a western coniferous forest.
UCRL- 53109. Lawrence Livermore National Laboratory, Livermore, California.
72 pp.
Kercher, J.R., M.C. Axelrod, and G.E. Bingham. 1980. Forecasting effects of SO2 pollu-
tion on growth and succession in western conifer forests. Iji: Effects of Air
Pollutants on Mediterranean and Temperate Forest Ecosystems, ed . P.R. Miller.
USOA Forest Service Report PSW-43, pp. 200-212.
King, D.A. 1986. A model for predicting the influence of moisture stress on crop losses
caused by ozone. Ecological Modelling (In Press).
King, D.A., J.R. Kercher, and G.W. Bingham. 1983. Modelling the effects of air pollu-
tants on soybean yield. Xn: Analysis of Ecological Systems: State-of-the Art
in Ecological Modelling, eds. W.K. Lauenroth, G.V. Skogerboe, and M. Flug.
Amsterdam: Elsevier Scientific Publishing, pp. 545-552.
Krupa, S. and D.W. Gardner. 1986. Agricultural Experimental Station Technical Bulletin.
University of Minnesota (In preparation).
Krupa, S. and R.N. Kickert. 1987. An Analysis of Numerical Models of Air Pollutant Expo-
sure and Vegetation Response. Prep for the Acid Deposition Research Program by
the Department of Plant Pathology, University of Minnesota, St. Paul,
Minnesota, U.S.A. and Consultant, Corvallis, Oregon, U.S.A. ADRP-B-1 0-87 .
113 pp
Larsen, R.l. and W.W. Heck. 1976. An air quality data analysis system for interrelating
effects, standards, and needed source reductions. Part 3. Vegetation injury.
Journal of the Air Pollution Control Association 26: 325-333.
Lefohn, A.S. and H.M. Benedict. 1982. Development of mathematical index that describes
ozone concentration, frequency and duration. Atmospheric Environment 16:
2529-2532.
113
Lockwood, J.L., J. A. Percich, and J.N.C. Maduewisi. 1977. Effect of leaf removal simulat-
ing pathogen-induced defoliation on soybean yields. Plant Disease Report 61:
458 -462.
Loehman, E. and T. Wilkinson. 1983. Ozone damage to field crops in Indiana. Station
Bull. No. 426. Agricultural Experimental Station, Purdue University, West
Lafayette, Indiana. 38 pp.
Luxmoore, R.J. 1980. Modeling pollutant uptake and effects on the soi 1 -plant -1 itter
system. I_n: Proceedings of the Symposium on Effects of Air Pollutants on
Mediterranean and lemperate Forest Ecosystems, ed . P.R. Miller. 1980 June
22-27; Riverside, California; USDA Forest Service Report PSW-43; pp. 174-180.
Male, L.M. 1982. An experimental method for predicting a plant yield response to pollu-
tion time series. Atmospheric Environment 1 6( g) : 2247-2252 .
Medeiros, W.H., P.D. Moskowitz, E.A. Coveney, and H.C. Thode, Jr. 1983. Oxidants and
acid precipitation: a method for identifying and modeling effects on United
States soybean yield. 83-2.4 Proceedings of the 76th Annual Meetings, Air
Pollution Control Association, Pittsburgh, Pennsylvania. 9 pp.
Miller, P.R., O.C. Taylor, and R.G. Wilhour. 1982. Oxidant air pollution effects on a
western coniferous forest ecosystem. US Environmental Protection Agency,
Environment Brief. EPA -600/D-82 -276 . 1 0 pp.
Mortensen, P. 1984. Modelling ion uptake in agricultural crops. RISO-M-2446. Riso
National Laboratory, Roskilde, Denmark. 33 pp.
Nosal, M. 1984. Atmosphere-biosphere interface: analytical design and a computerized
regression model for Lodgepole Pine response to chronic, atmospheric SO2
exposure. RMD Report 83/26 and 83/27. Research Management Division, Alberta
Environment, Edmonton, Alberta. 97 pp.
Nosal, M. 1983. Atmosphere - bi osphere interface: probability analysis and an experimental
design for studies of air pol 1 utant -i nduced plant response. RMD Report 83/25.
Research Management Division, Alberta Environment, Edmonton, Alberta. 98 pp.
Nosal, M. and S.V. Krupa. 1986. Numerical methodology in the risk assessment of air
pollutant induced ecological effects. 86-92.1. Proceedings of the 79th Annual
Meetings, Air Pollution Control Association, Pittsburgh, Pennsylvania. 16 pp.
Oshima, R.J., M.P. Poe, P.K. Braegelmann, D.W. Balding, and V. van Way. 1976. Ozone
dosage-crop loss function for alfalfa: a standardized method for assessing crop
losses from air pollutants. Journal of the Air Pollution Control Association
26: 861-865.
Pratt, G.C. and S.V. Krupa. 1985. Aerosol chemistry in Minnesota and Wisconsin and its
relation to rainfall chemistry. Atmospheric Environment 19: 961-971.
Pratt, G.C, R.C. Hendrickson, B.I. Chevone, D.A. Chri stopherson , M.V. O'Brien, and S.V.
Krupa. 1983. Ozone and oxides of nitrogen in the rural upper -mi dwestern
U.S.A. Atmospheric Environment 17: 2013-2023.
Rowe, R.D. and L.G. Chestnut. 1985. Economic assessment of the effects of air pollution
on agricultural crops in the San Joaquin Valley. Journal of the Air Pollution
Control Association 35: 728-734.
Runeckles, V.C. 1974. Dosage of air pollutants and damage to vegetation. Environmental
Conservation 1 : 305-308.
Sisterson, D.L., B.E. Wurfel , and B.M. Lesht. 1985. Chemical differences between event
and weekly precipitation samples in northeastern Illinois. Atmospheric
Environment 19: 1453-1469.
Stevens, T.H. and T.W. Hazelton. 1976. Sulfur dioxide pollution and crop damage in the
Four Corners region: a simulation analysis. New Mexico Agricultural Experi-
mental Station Bull. 647. Las Cruces, New Mexico.
114
Stevens, R.K. T.G. Dzubay, G. Russwurm, and D. Rickel. 1978. Sampling and analysis of
atmospheric sulfates and related species. In: Sulfur in the Atmosphere, eds.
R.B. Husar, J. P. Lodge, Jr., and D.J. Moore. New York: Pergamon Press, pp.
55-68.
Teng, P.S. and R.E. Gaunt. 1980. Modelling systems of disease and yield loss in cereals.
Agricultural Systems 6: 131-154.
West, D.C., S.B. McLaughlin, and H.H. Shugart. 1980. Simulated forest response to chronic
air pollution stress. Journal of Environmental Quality 9: 43-49.
Westman, W.E. 1979. Oxidant effects on California coastal sage scrub. Science 205:
1001-1003.
Woodmansee, G. 1974. Glossary of systems ecology terms. Technical Report No. 256. US
International Biological Program, Natural Resource Ecology Laboratory, Colorado
State University, Fort Collins, Colorado. 32 pp.
115
5. EFFECTS OF ACIDIC DEPOSITION ON SOILS
The acidification of soils as a result of acidic deposition and the potential
effects on terrestrial ecosystems have been topics of considerable concern in the world
in recent years. The potential for, and documentation of, such effects have been
focussed in studies on the acidic soils of Scandinavia, northern Europe, northeastern
United States, and southeastern Canada. Concern relates primarily to the potential for
reductions in forest productivity as a result of acidic deposition and on the potential
for acidification of lakes and streams with subsequent reductions in aquatic productivity
via soil throughflow and/or overland runoff. More recently, and particularly in Alberta,
the potential for adverse agricultural effects which may result from acidic deposition
have begun to be addressed.
The following narrative briefly outlines the processes that take place in soils
during acidification, the soil buffering systems, and possible mechanisms of the process.
The latter part of this discussion will then place these acidifying processes in a world
perspective and look at the implications of such effects on soil types in general.
5.1 ACID-BASE SYSTEM IN SOILS
The Bronsted-Lowry concept of acids and bases is commonly applied to soil acids
and bases (van Breemen et al. 1983). Soil is composed of many conjugate acid-base
pairs. Examples of such acid-base pairs are: HaO"*" - H2O; A1(0H)2"*" - A1(0H)3; H2CO3 -
HCOa"; and NHn"^ - NHa.
The strength of acid-base systems in soils is measured by the proton
dissociation constant (K^^):
HA + H2O = HaO"^ + A"; K^^ = (HaO"^) (A~)/(HA) [1]
pH = pK^^ + log [(A")/(HA)] [2]
When the acid system is strong, the dissociation of protons is increased and the value
of K^^ is high; the pK for an acid is the pH value at 50% dissociation where (A ) =
(HA).
Soil acidity is best considered in terms of intensity and capacity factors.
Capacity factors are a function of the size or quantity of the system and are directly
influenced by addition or depletion of protons. Capacity factors are characterized by
the amount of base needed to titrate the soil to a set end point. Intensity factors
such as pH are a function of the chemical properties of the system and are independent
of its quantity (van Breemen et al. 1983). Bache (1980) stated that there are three
parameters included in defining soil acidity: the total acidity, the degree or intensity
of acidity, and the buffer capacity or the manner in which the degree of acidity varies
with total acidity.
5.2 SOIL REACTIONS
The degree of acidity usually represented by pH values is related to the number
of free hydrogen ions in solution, which probably exist as the hydrated forms of HaO"*" and
HtOs^. However, soils are generally mixtures of porous charged solids and humus with
116
little solution, which makes it impossible to define theoretically or determine experi-
mentally a unique pH value (Bache et al. 1979a). Measured soil pH values have been found
to depend on the following variables as defined by Jackson (1958): drying of the soil,
soil-to-water ratio, CO2 concentration in equilibrium with soil suspension, and solu-
tion electrolyte concentration. Because of the variance in soil pH depending on its
environmental state, some researchers have suggested alternate methods for its mea-
surement; for example, measurement of the lime potential (Schofield and Taylor 1955;
Bache 1979a) or the addition of a standard method of pH measurement made at unique
soil-to-water ratios and salt concentrations (Peech 1965).
Currently, soil pH measurements are mainly conducted potentiometrical ly in
dilute electrolyte suspensions. However, such measurements can cause problems because
the pH values obtained can vary depending on the position of the calomel electrode in
the suspension (Black 1968; Thomas and Hargrove 1984). The suspension effect can be
eliminated by suspending the soil in an electrolyte solution (CaCl2 or KCl ) of ionic
strength greater than 0.005M (Bache 1979a).
Lime potential or the pH of soil measured in a supernatant of O.OIM CaCl2 was
proposed by Schofield and Taylor (1955) to circumvent the problems of measurement. They
expressed their results as pH-l/2pCa or "lime potential", where pCa is the negative log
2+
of the Ca activity in solution which is equivalent to the mean activity of Ca(0H)2.
Lime potential is considered independent of electrolyte concentration at least for those
encountered in soils (Thomas and Hargrove 1984).
In calcareous soils the activities of H^ and Ca^^ are determined by the
following reactions:
[CO2] soln - k^ p CO2 [3]
n
H2O (- [C02]soln - H2C03 = [H*"] + [HCOa']
^ 2[h''] + [COa^'l [4]
CaCOa = [Ca^^] + [COa^"] [5]
The lime potential of soil thus depends on the partial pressure of CO2 and the temper-
ature of the soil (Talibudeen 1981). In non -cal careous soils, the activities of H^
and Ca^^ ions are determined largely by the following reactions (Thomas and Hargrove
1984; Reuss 1985):
A1(0H)3 + 3 H"*" = Al""^ + 3 H2O [6]
K = (A1^')/(H^)^ [7]
a
3 CaX I- 2A1^'^ = 2 AIX + 3 Ca"""^ [8]
Kg = [(AIX)" . (Ca''')']/[(CaX)' • (Al'')'] [9]
117
Combining the above equations for K and K , and rearranging them gives the
a g
f ol lowing:
[(Ka)'/'Kg'/^CaX)'/^]
(Ca^")^/V(H") = ^-j- [10]
Taking the negative logarithm of both sides gives the following final formula:
pH - 1/2 pCa = [11]
where K^^ - lime potential, which is equivalent to the negative logarithm of the right
hand side of the equation. This equation 11 states that the lime potential is propor-
tional to the Al -H proportionality constant (K,), the exchange constant (K ), and
a y
the exchange sites occupied by Al and Ca.
The usefulness of the lime potential concept is based on the premise that soil
reaction is a product of the interaction and balance of ions and other mobile ions,
whose relationship is best expressed as the activity relative to the reduced ratio
of other cations (Bache 1979a). The other cations include Ca^^, Mg^^, and Na^ in cal-
careous soils plus Al^^ in acidic soils.
Buffering in soils is chiefly due to colloidal organic and inorganic materials.
Buffering capacity may vary with pH and may also be time dependent as dissolution
kinetics vary with the composition of the soil. The inorganic colloidal soil complex
functions as a slightly ionized acid or as a slightly ionized salt of a weak acid.
Clays act as weak acids due to exchangeable aluminum which hydrolyzes and exhibits
different forms varying with pH as shown below:
A13+ < — A1(0H)2+ < Alp (OH) ^ (3n-m)+ <_.__ A1(0H)3 <- A1(0H)4
mono-nuclear poly-nuclear solid [12]
ions ions
approximate pH
3.5 5 6.5 8
There are four major soil buffering systems all of which are pH dependent.
They are: aluminum buffering (approximately pH 4), iron buffering (pH range less than 3
to 3.5), carbonate buffering (pH range 6.5 to 8.3), and organic buffering (pH range 4.5
to 6.5) (Ulrich 1980) .
The basic formulas and reaction types for each type of buffering system are as follows:
Aluminum: AIOOH.H2O + xH Al(OH) ^'^^+ XH2O [13]
Iron: Fe(0H)3 + xH"^ = Fe(OH)^tj^ + XH2O [14]
[15]
Organic: R - COOH + H2O = R-COO~ + H30'^ [16]
118
In the normal pH range of soils, the organic soil colloids can act as a buffer-
ing system. Carboxylic (-COOH) groups are chiefly responsible for acidity in organic
soils within the pH range of 3 to 7. Phenolic -OH groups contribute acidity between
pH 8 and 12. Between pH 7 and 8, acidity is contributed by both carboxylic and
phenolic-OH groups plus x-NHa groups (Gessa 1979). The range of the organic buffering
system spans all of the above pH values and is therefore active in most soils.
5.3 TOTAL SOIL ACIDITY
Many approaches have been used to characterize the components of total soil
acidity (Bohn et al. 1979). Two generally recognized components are exchangeable and
non-exchangeable acidity. These are differentiated on the basis of the experimental
method used to measure the particular fraction.
Total acidity is defined as the amount of base required to bring the soil to a
pre-determined pH value under standardized conditions (Bache 1980). It is the combined
total of free h"*" plus the undissociated forms of acidity in the soil (Holowaychuk and
Lindsay 1982). It is, therefore, the sum of both exchangeable and non-exchangeable
acidity. The titration parameters and methodology used in determining total acidity
must be specified for this type of measurement to be useful. Soil titrations are
particularly susceptible to experimental variations such as method of stirring, time of
treatment, and period between base additions.
Exchangeable acidity is only detectable in soils having a pH value lower than
5.5. In such soils, exchangeable aluminum primarily constitutes the acidity below pH 4,
although exchangeable hydrogen contributed by aluminum hydrolysis is also measured by
this technique (Bohn et al . 1979; Thomas and Hargrove 1984). Exchangeable acidity,
which comprises a portion of the total acidity, is estimated by leaching or extraction
of soils with a IM solution of a neutral salt and titrating the extract with base (Bache
1980).
The difference between total acidity and exchangeable acidity is the non-
exchangeable acidity. This form of acidity constitutes the major part of total acidity
in soils and exists mainly in undissociated forms (Bache 1979a). There are three main
derivations of non-exchangeable acidity: neutralization of hydroxyl-Al polymers at soil
surfaces; neutralization of hydrogen ions from organic functional groups; and displace-
ment of adsorbed anions (Bohn et al . 1979). These processes occur mainly within the pH
range of 5.5 to 7.0 but can also occur at higher pH values (Bohn et al. 1979).
5.4 CATION EXCHANGE AND SOIL ACIDITY
The negative charges found on the colloidal clay and humus fractions of the
soil matrix give rise to the phenomenon of cation exchange. The cation exchange capacity
of a soil is composed of a constant permanent charge and a variable pH-dependent charge
(Bache 1979b).
The permanent charge is independent of soil pH and is generally constant for
the pH range of most soils. This charge is derived from the isomorphic substitution of
Al^"^ for Si^"*" in the tetrahedral layers, and of Mg^"*" or Fe'''*' for Al'"^ in the octohedral
layers of clay-sized layers of silicate (Bache 1979b). The result of such substitution
is that hydroxyl and 0^~ charges in the clays become unbalanced and acquire a net
negative charge.
119
The source of the variable charge in soils is the pH-dependent dissociation of
functional groups on the surfaces of soil solids. These groups include the following:
hydroxyl, carboxyl, phenolic, and amine in soil humus, and the aluminol and silanol
groups on the crystal edges of layer silicates and the surfaces of al uminosi 1 icate gels
(Bohn et al. 1979). Variable charges are also derived from blocking of negative charges
by adsorbed hydroxyl aluminum cations (Bache 1979b).
5.5 BASE SATURATION
Base saturation is defined as the ratio of basic exchangeable cations to the
total cation exchange capacity (CEC) of the soil. These base cations include Ca, Mg,
Na, K, and NH*. The degree of base saturation (BS) is formulated as follows:
Base Saturation = [(Ca + Mg + Na + K + NH4)/CEC]100 [17]
The degree of base saturation is dependent on the pH at which the CEC measurement was
made; they are inversely related.
A two step ion exchange process is used to obtain CEC values. In the first
step, soil samples are leached with salts to extract exchangeable cations and saturate
exchange sites with an indexed cation; in the second step the indexed cation is leached.
CEC is then calculated from the total concentration of the indexed cation in the
leachate.
There are three main methods of obtaining CEC which vary depending on the
starting pH of the soil, and unfortunately, each gives a different value for CEC but the
same value for base saturation. Bache (1979b) stated that the relationship between CEC
and exchangeable ions can be represented as follows:
CEC = exchangeable + exchange acidity + hydrolytic [18]
base cations (acid cations) acidity
As the pH of the extracting solution increases, the amount of exchangeable base
cations stays constant but CEC increases due to increased acidity. This results in a
decrease in base saturation as the pH rises. The strong interdependence of base satura-
tion on pH of the extracting solution used to determine CEC makes base saturation an
unreliable measure of soil saturation with base cations or unsaturation with acid
cations. The most realistic picture of any losses in base saturation due to acidifica-
tion should be obtained with measurements taken at the native soil pH.
5.6 NATURAL ACIDIFICATION OF SOILS
5.6.1 Acidification in Soil Genesis
Soils and their composition are considered to be a function of many different
factors such as parent material, climate, biota, topography, and time (Jenny 1941).
Parent material, topography and time are passive factors in soil development, whereas
climate and biological action are considered to be active factors which drive soil
processes.
120
Acidification is an increase in the total acidity of the soil and a reduction
in its pH. Over time this process causes a transformation of the chemical conditions of
the soil with respect to H^. Climatic influences on soil acidification result from
the effects of temperature and water on the weathering of material and the types and
levels of activity of soil biota. Acidic deposition on soils results in bicarbonate
weathering as the acid strips CaCOa from the soil. Other processes that also accom-
pany this acidification are hydrolysis, hydration, and carbonation. These processes
cause the decomposition of soil mineral constituents (Tabatabai 1985). Soil acidification
is enhanced in high rainfall areas, or as the duration of these acidifying processes
increases .
Decomposition of vegetation as a result of biological action in the uppermost
soil layers can result in the release of both bases and acids (inorganic and organic).
Production of acids is particularly high in the biologically active surface layers of
forest soils (i.e., litter zone). In comparison with grassland soils, forests are much
less efficient in recycling and retaining nutrients. In part, this is related to the
moisture regime of the ecosystem. Grasslands in general are drier and this results in a
lower leaching rate and in the retention of the primary buffering cations. Over time,
forest soils will become more acidic under natural conditions with acidic deposition
causing a possible acceleration of this natural effect. Changes in the pH of the litter
layer also directly affect the types of biota living there, and hence, the decomposition
processes that occur. This aspect, however, will be discussed in detail in a later
section of this overview and is only mentioned here.
5.6.2 Natural Sources of Soil Acidity
There are various sources of hydronium ions responsible for soil acidity. The
various sources, sinks, and pathways of soil acidity are shown in Figure 2 (Krug and
Frink 1983a).
5,6.2.1 Organic Matter. The organic layer of soils consists of live organisms (plant
and animal) and their undecomposed , partly decomposed or transformed remains. Humus is
part of this layer and is composed of the transformed remains of vegetation or animal
matter. Humus includes humic substances which are differentiated on the basis of pH
dependent solubility into fulvic, humic, and humin fractions, and non-humic substances
from the following classes: carbohydrates, proteins, lipids, and organic acids (Paul
1970; Oades and Ladd 1977).
In neutral and alkaline soils, most of the organic matter is in the humus form
with large proportions in humic and humin fractions. These materials have high concen-
trations of carboxyl, phenol hydroxyl, and other functional groups which dissociate to
produce hydrogen ions. These groups weather and decompose soil minerals by complexing
with and dissolving metals. Calcium is selectively adsorbed by humus material and the
Ca-humic complexes may counteract acidification (Wiklander 1979). In acidic soils,
humic and fulvic acids are present mainly as iron and aluminum complexes (Thomas and
Hargrove 1984). With the gradual decomposition of organic matter in such soils, Al and
Fe are released and can contribute to overall acidity by hydrolysis. However, in acidic
soils most of the acid produced is lost from the system as H2O or CO2 following
121
O
Acid Rain
H2SO4 HNO3 H2O + CO2
H2O + CO2
Ca+2Mg*-2K+Na+
©
Complete oxidation
Litter
©
Formation of humic residue by partial oxidation
Rn-C- H
Rn- C
Rn-i -C • • • + — *^ H2CO3
N.S.P
©
r
©
OH
O ©
OH
0
H+
Rn-C Rn-1 -C
H+
HCO3-
H+
I ^ I
J
Mineral weathering
©
Ca+2Mg+2K+Na+
rr ^
Polyvalent cations (3c)
Biological uptake
©
Secondary minerals.
©
©
Acid export
Cation exports: Ca*2Mg*2K+Na+ Aha
Anion exports: S04"2N03"HC03-RCOO-Cr
H*
Figure 2. Major sources and sinks of acidity in soil (Source: Krug and
Frink 1983a).
Acid rain (1) is a source of acidity, and its composition may be altered
before reaching the soil (la). Although biological processes (2) are net
sources of acidity, this obscures the fact that they serve as a substan-
tial sink in acid soils through production of weak organic acids (2b) with
ultimate conversion to CO2 and H2O (2c). Mineral acids (2a) can be cycled
rather tightly with some S and N lost to the atmosphere by microbial
activity, and some S and P can be converted to essentially insoluble
secondary minerals. Weathering of minerals (3) generally consumes acid
in excess of cation export (3d), as secondary minerals (3b) and hydrolysis
products of aluminum, iron, manganese (3c) accumulate in soil. Aggrading
vegetation causes net cation uptake (3a), and contributes to acidifica-
tion. Rain less acidic than the soil solution promotes acidification by
hydrolysis (3c). The electrical charges exported by cations (3d) and adds
(4) are balanced principally by anions shown at the bottom of the figure.
122
decomposition, rather than contributing to weathering or being translocated through
runoff .
Organic acids can have an acidifying influence on soils. For example, the
litter layer in coniferous forests is very acidic and produces an acidic humus. The
soluble organic substances, including fulvic acids of this soil horizon, are leached and
cause strong acidification and weathering which eventually lead to the formation of
podzols with low base saturation (Wiklander 1979). Podzol A horizons are developed by
leaching while illuvial B horizons result from accumulation of Fe, Al , and humic
materials. Bases do not moderate the acidifying influence of this process primarily
because they are maintained in the humus layer by biological cycling and do not leach
out.
Litter from deciduous hardwood forests is higher in base content than that of
coniferous forests. Consequently, hardwood litter tends to increase the alkalinity and
the overall buffering capacity of the humus layer, thus making it more resistant to
acidifying processes. Grasses are even more effective than trees in reducing acidifi-
cation since they maintain high base contents and resist leaching of the major nutrient
salts by having highly efficient recycling capabilities (Tabatabai 1985).
Ammonium represents another source of potential acidity in soils. There are
three primary sources for ammonium in soils: decomposition of vegetational matter by
soil microbes; fertilizer applications; and atmospheric deposition. In soils, the
ammonium ions are microbially oxidized to form nitric acid. Theoretically, 1 mole of
ammonium would produce 2 moles of hydrogen ions following nitrification. In practice,
however, direct uptake by plants, volatilization, denitrif ication processes, and the
high ratio of nitrogen uptake to excess base uptake by plants reduces the net effect on
soil acidity caused by ammonium additions.
The oxidation of sulphur compounds such as S, FeS, FeSa, and H2S also generates
acidity. Sulphur, in the sulphate form, is taken up by plants and incorporated into
plant materials in its various reduced organic forms. Because of this action, although
the original oxidation of sulphur compounds generates soil acidity, no net change in
acidity occurs as the sulphur goes through this reduction cycle (Holowaychuk and Lindsay
1982) .
In anaerobic soils, sulphate is reduced to sulphide, leading to H2S or FeSa
(pyrite). If these soils then become aerobic, acidity is produced by oxidation. Highly
acidic soils with pH values as low as 2 have been created by this type of process in,
for example, floodplain areas (Thomas and Hargrove 1984). Conversely, waterlogging or
flooding of soils leading to anaerobic conditions can reverse this process, thereby
counteracting the acidification process (Wiklander 1979).
Nutrient uptake by plants can also have a significant effect on soil acidity.
Cation exchange, whereby nutrient cations are adsorbed by root surfaces, causes the
desorption of hydrogen ions by the root, resulting in base cation replacement in the
soil and a net change in soil acidity. Wiklander (1979) stated that this action likely
causes a depressive effect on base saturation in fertile and cultivated soils but found
the exact magnitude of change difficult to quantify. The form of nitrogen taken up by
plants was also found by Ulrich (1980) to have a bearing on soil acidity. If, for
123
example, nitrogen was taken up as NOa" no acidity was produced, but if nitrogen was
taken up as NH4"^, up to 4 kmol (H"^) ha~^ could be produced.
In summary, the above discussion indicates that soil acidification is a natural
process which produces an excess of h"^ over time. The mechanisms and intensity of
acidification are dependent on the soil forming processes. Table 29, adapted from
van Breeman et al. (1983, 1984) summarizes the hydrogen ion producing and consuming
processes in soils.
5.6.2.2 Leaching and Weathering. Leaching is the downward movement of dissolved
substances through soils as a result of water movement. Substances prone to leaching
include soluble salts, bases, silicon, and various forms of organic material. Leaching
mechanisms involve many variables such as the nature of the soil medium, activities of
microorganisms, formation of complex ions, surface charge and exchange properties of
soil particles, partial pressure of CO2, and porosity and hydraulic conductivity of
the soil (Finkl 1979). The degree and type of leaching differs under acid, neutral, and
alkaline conditions and under oxidizing versus reducing conditions.
The effect of leaching on soil chemistry varies according to the substances
being acted upon. Readily soluble substances such as salts usually percolate with water
from upper to lower soil horizons. On the other hand, the alkali and alkaline elements
either in mineral or exchangeable form are mobilized by hydrolysis by dilute acids such
as carbonic acid. The kind and extent of mobilization is governed by exchange reactions
with cations in the percolating waters (Holowaychuk and Lindsay 1982). The bonding
energy of cations on exchange sites and in soil solution influences the extent to which
different cations are displaced or adsorbed. At a pH of 5.5 or higher, hydrogen ions
have high bonding energy and are very efficient in displacing bases.
Weathering is accelerated under high leaching rates. In addition to mobiliza-
tion of alkali and alkaline earth elements during hydrolysis, silicon can also be
released and leached, mainly as silicic acid. Because leaching reduces the concentration
of soluble weathering products, conditions conducive to the continuation of weathering
are produced. These processes not only contribute to loss of bases but, in the case of
primary aluminosi licate minerals, to an increase in the aluminum content of the weather-
ing residues (Holowaychuk and Lindsay 1982). Aluminum and iron form complexes and
chelates with acid radicals or functional groups of humic materials. In these forms,
iron and aluminum are subject to leaching but not to as large a degree as the alkali and
alkaline earth elements or silicon. The net result of these processes is that bases are
depleted from soils, and acidity is increased due to higher proportions of hydronium
ions and aluminum and to an accompanying decrease in base saturation of the exchange
complex.
In acidic soils, the soil solution tends to be enriched with carbon dioxide and
humic materials. The alkali and alkaline earth elements are readily leached, while
humus and silicon are leached to a lesser degree. Iron and aluminum, however, are
removed in small amounts. In neutral soils, the alkali and alkaline earth elements are
also leached but tend to be deposited in deeper horizons as carbonates and sulphates.
Leaching of aluminum, iron, silicon, and humus is negligible in neutral soils. In
alkaline soils, leaching is restricted because these types of soils mainly occur in arid
124
l/l
c
o
«/i
a>
c
"O
c
O)
o
ro
c
i/i
X
(O
O)
+J
tz
1/1
o
M-
JC
1/1
o
c
O
o
i/i
o
o
X
O)
c
ro
Ol
<u
ro
4-)
o
o
(O
<a
c
c
i-
E
iZ
o
o
o.
c
O
o
o
+->
■M
O)
c
o
+->
ro
C2.
■M
o
c
ro
O
J-
ro
o
l/l
cn l/l
o;
\
o
*i
rsi
+->
a
c -t-J
3
E
3
O)
ro
+-*
o
•t— c
3
■o
in
05
ro
i- <D
CD
ro
c
ro
to
c
•(-'
<u c
l/l
C
o
C
s_
E
o
o
o
S_
O
C
o
<u
■H
3
+J CL
CL)
4-
o
<a
c
1/1
o
■o
ro E
>
TD
c
S-
1/1
s_
<u
<X) o
>)
c
o
o
<C
cn
31
<:
+->
o
, ^
, — ,
. — .
, ^
^ — .
, ^
, ^
S-
CO
00
ai
Q-
.
^-^
+
cz
3:
o
+
■(->
3
+
CM
^ — .
3=
+
o
C\i
+ rs
ro
D-
3:
1/1
ro
CM
+
+
1 O
3:
+ <N 00
+
+
+
3: 'J CM
3:
3:
o
+
CM O
c
E
1
l/l
+
00 +
CM
rN
o
+
+
OO
O"
+ CM
+
m
+ PI
+
O
cn
H
31
O
n ^
+
00
+
+
+
Z
^ m 3:
l/l
1
31
1/1
(U
+
31
+
3: ^ o
E
+
0}
+
+
1 + +
•X.
+
CM
II 3:
o 3: — '
o
CM
+
CO
«/)
CL
+ 3:
1 <N :e
o
+
CM
o a>
c
rN
OD
I/I
E
+ I
o
CD U_
\
1
+
cn
0)
03
+
o o + +
cr
CM
<N +
CN
+
3:
3:
u
X
+
Q- 00
ro
1
3:
s:
E
O
c
o
o
UJ
o
<X 1 1
II
<N +
II
c
CM
Q.
i-
T3 <N
:n + p> <n
CM ro
II o
CM
rN
+
o.
i-
c i:
O O
1
+ o o
.\ O
o
\
O
JC
0)
+ M Z Z
o
+
c
:e 00 <N
1— Z
rM O 31
Z
u
ro
o>
+->
O) +
3:
c
CM :r
CO
IE M
+
X
c
fO
3: + +
o
II
+
3: CM
X
LU
+->
3
+ <N
o o
+
+ +
II
CvJ \
«•
0)
*E
» H 31
c
o
rN
\ CM r-~
o
o
o
c:
Q£ • O O
1
<t
1 1
O O
in \
(N
Cs-
s:
1/1
+ •
o: • •
+
m +
«N 3: m
^ (N
m +
3:
3=
:c
o
ro
c
(O
II en
II O::
o i:
II
« O
X 3:
+
c
E
ro
II
rM
E
O
J-
II
o
+ C_)
O + Z
CM
+ rN
CM
CJ
3:
0)
u
■o
o-o II
O II II
^ 31 +
OO
1 +
rN O
\
+
CM
cu
\
«N O
3:
cr
<M II
uo
O rx
i_
-o
e
»o :^
3: <N o (M
+ li 1
ro
II O
fM
O 'd-
^ — .
X
+
CQ
c
II cn O
31 M O
CO
o
O
•d- \
+
00
UJ
lO
(O
+ CVJ ^ CM
\
1 o •
rx CM O
■ — CM Ln
c
s_
jc—i a:
CO
3: <N q:
c
O \ CVI
\ CM
:r
ro
o>
o •
t + + +
o ^
o
CM 1—
t— \
cn
c
O
X
>
c
cr
a: +
o
+
II
+
in
+ +
fO
+
O-
UJ
Q. « «1 <N
II +
c
+ +
+
+
E
CN
+
c=
u
+ + +
<N o o
31
3:
<t
+
+ + oi
31
3
:e t/) z z
00
O <M O
OO <N «
X
rN 00 00
+
o
+
■o
+
+ + 3:
M O •
c:
<M o 3:
O rN
O; (U 0)
z
(O
c
o
00
o
t:
s: z z
C£ Q£ Q£
cc
:r <_) q;
31
31 00 Z
z z
U- U- LL.
C_)
00
0)
o.
xa
c
ro
c
<u
o
+->
o
l/l
D1
■fj
1/1
c
C
-t-J
o
c.
o
ro
Q.
ro
l-
a>
o
l/l
<4-
o
i-
c
Q.
1/1
c:
O
o
i/i
-M
ro
c
X
C31
cu
re
o
01
o
0)
a>
o
(_)
+->
(_)
CO
c
"O
l-
o
3
o
c
E
o
>
4-
O.
<4—
*[I
«0
3
Q.
c
o
c
o
a>
0)
a;
o
JC
e
CJI
sz
i
05
c:
c
4->
o
o
c
+-»
C
u
o
+->
o
ro
S-
ro
3
ro
l/l
0)
01 00
■o
E
</J
(_)
i_
+j
N
■H
c
3
c ■»->
>1
<-)
o
3
ro
10
o
•r- C
^
X
i-
•o
O) i>o
i- 0)
\
a;
<4-
O
Q.
ro
u
+J
l/l C
<D C
cz
s_
"i
s_
o
s- o
^ o
o
c
XJ
o>
Q.
O
(U
T3
Q) -r-
+J Q.
o
<u
CVJ
E
1/1
c
l/l
> +->
ro E
■H
+J
C=
+->
1/1
X
CD ro
<D O
ro
c
Q.
a>
o
<
<:
^
a
o
oe. o
3 <-)
(_)
<t
ro
+->
T3
o
<C
«
C\J
CO
r-
00
cn
a.
H
125
regions. The effects of leaching on acidification are thus greatest in those soils
which are already acidic, and it would appear that any additional acidic inputs would
accelerate acidification and leaching of base ions.
5.7 INFLUENCES OF SOIL ACIDITY AND ACIDIFICATION ON SOIL PROPERTIES
5.7.1 Organic Hatter
The physical and chemical forms of organic matter differ among soils and, as a
result, a number of organic horizons are widely recognized (Buol et al . 1980). For the
purposes of this discussion, the following broad classifications are recognized:
1. organic matter of Chernozemic soils and surface layers of cultivated
soils; along with the melanic organo-mineral horizons of deciduous forest
soils, this type is commonly referred to as mull ;
2. the L-F-H layers, in different combinations, of forest soils; this type is
equivalent to mor as described by Buol et al. (1980);
3. organic matter of Solonetzic B horizons, which occurs as dark coatings on
the surfaces of peds;
4. organic matter of Podzolic B horizons occurring as dark reddish brown to
black, soft, weakly granular layers;
5. organic matter of Organic soils, or peat, which can occur in various
degrees of decomposition.
The organic fraction of soils accounts for the major portion of pH dependent
CEC (Bache 1976), and from 25% to 90% of the total CEC of surface horizons of mineral
soils (Stevenson 1982). The relationship of CEC to pH for various organic soil types is
shown in Figure 3. From this graph it can be seen that CEC and pH are directly related
and that a decrease in one results in a decrease in the other. The CEC-pH relationship
varies, however, according to the type of organic soil. For example. Figure 3 shows
that the CEC of organic soils is much more strongly influenced by pH than CEC of clay.
The negative charges associated with humus are also dependent on pH. The covalently
bonded hydrogen of carboxyl and phenol groups is not dissociated at low pH values.
Hydrogen dissociation occurs with increasing pH, and negative charges on the colloids
develop as a result. For these reasons, the amount of humus in a soil, even if only in
a small proportion, can significantly contribute to the total CEC value. Many of the
exchange sites of organic matter in mineral soils may be blocked because of the formation
of organo-clay and organo-metal complexes. The acidification of soil would have the
effect of reducing effective CEC and displacing base cations which may then be removed
by leaching. One possible remedy to this problem would be to lime the soil, which would
effectively increase the CEC by disrupting the organo-mineral complexes and by causing
the dissociation of carboxyl and phenolic groups (Stevenson 1982).
Low soil pH (<4.0) has been reported by Jenkinson (1981) to decrease the rate
of organic decomposition of rye grass. This decrease was thought to be caused by a
decrease in the numbers of microbial species capable of functioning at low pH values.
The acid tolerance of fungal species responsible for decomposition in forest litter.
126
Figure 3. Reduction in cation exchange capacity of organic matter and
clay with decrease in soil pH.
A - Organic matter fraction in Ap horizons (Helling et al.
1964)
B - Humus of forest soils (Kalisz and Stone 1980)
C - Clay fraction in Ap horizons (Helling et al. 1964)
127
makes that system relatively non-responsive to changes in pH although some slight effects
are apparent (Abrahamsen et al. 1980). McFee (1982) has listed the suspected results of
acidification on organic decomposition as including decreased rates of carbon minerali-
zation as a result of lowered pH or associated heavy metal toxicity, shifts in the
microbial community structure away from bacteria towards acid resistant fungal dominated
communities, and decreased ammonif ication and nitrification.
Acidification can also be caused by application of industrial fertilizers.
Barratt (1970) found that the soil morphology and humus form was changed as a result of
the application of ammonium sulphate fertilizers. In this study, the pH of soils in hay
fields dropped (pH 4.5 to 3.9) and the organic horizons of the soil developed the
properties of a mor soil. These changes were accompanied by a reduction in decomposition
rates and most notably, earthworm activity. Other types of fertilizers did not have
these effects on increased acidity, and mull morphology was maintained. In another
study, Goh et al. (1986) reported on the physico-chemical effects of the long-term use
of urea and ammonium based fertilizers on a Black Chernozemic soil in Alberta. The main
effects detected by this study were that the soil exhibited a lowered pH, altered balance
among Al and basic cations, exchangeable acidity, and titratable acidity. Although Goh
et al. (1986) found that the organic content of the soil increased with fertilization,
examination of its microstructure indicated that deterioration had occurred, probably as
a result of aggregation.
Schnitzer (1980), in his review on the effects of low pH on the chemical
structure and reactions of humic substances, indicated that under moderately acidic
conditions, fulvic acids are soluble and mobile while humic acids become aggregated and
immobile. These differences in reaction characteristics are thought to be caused by the
structural configuration of humic acids. Humic acids are thought to exist as randomly
coiled polymers in solution and are most tightly coiled and cross-linked in the centre
(Stevenson 1982). Saturation with protons or polyvalent cations, or dispersal in high
electrolyte solutions causes the coil structure of humic acids to shrink. Shrinkage is
caused by association of ionizable groups and cross-linking of polymers through inter-
action of functional groups and polyvalent cations. In acidic soils, the most important
species causing humic acids to undergo structural shrinkage is Al^^, while in
neutral or alkaline soils Ca^^ is the dominant agent. This cross-linkage results in
smaller, more dense and rigid particles which are thought to be more stable and resistant
to biological breakdown (Turchenek et al . 1987). This theory has not been substantiated
and at present the structural make-up of humic substances in soils is not known. How-
ever, the implications are that increased acidity will cause the aforementioned
structural changes with the net result being a decrease in the rates of organic
decomposition.
In soils where fulvic acids are the major organic component, such as Podzols,
acidification can result in the loss of organic matter due to solubilization and leaching
(Schnitzer 1980). Exceptions to this rule occur within the pH range 2 to 3 when fulvic
acids may be adsorbed to mineral surfaces, or in clay soils where interlayer adsorption
may increase as the clay minerals expand. Increased acidity and fulvic acid leaching
result in acceleration of soil weathering which in turn causes a reduction in biological
activity and a general lowering of soil productivity (Schnitzer 1980). These effects
are extremely rare in soils except under extremely acidic conditions.
128
The following publications summarize the types of organic matter and mineral
interactions and product formation mechanisms under various soil environments, and
should be consulted if more detail on this topic is desired: Schnitzer 1978; Burchill
et al . 1981; and Stevenson 1982. The role of organic matter and its interactions with
soil mineral constituents in aggregation of soils has been reviewed by Tisdall and Oades
(1982), and by Oades (1984).
Another effect of increased soil acidity on fulvic acid colloids could be an
increased tendency for dispersion and translocation causing the acceleration of the
podzolization and lessivage processes. Such induced dispersion may accelerate the
development of surface crusts and hard setting properties in certain types of tilled
soil. These processes can occur in soils with no free CaCOa and little solubilized
aluminum at a pH of about 6 (Oades 1984). These types of conditions are reasonably
prevalent on many of the tilled Luvisolic soils of western and central Alberta (Turchenek
et al. 1987).
Clays and organic materials are formed of polyanions which can be bridged by
polyvalent cations such as Ca, Mg, Al , and Fe. The minor elements such as Mn, Zn, and
Cu also contribute to bridging. The cation bridges can be disrupted by treatment of
soil with complexing agents and acids. The effect of such disruption is to cause soil
aggregates to become destabilized (Hamblin and Greenland 1977).
Clay-humus interactions can be produced by the introduction of lime which in
effect causes the cation bridge mechanism to be activated. The results of such additions
have been shown to stabilize soil structure and improve tilth in tilled Luvisols of
northeastern Alberta (Hoyt et al . 1981). Thus, in addition to its acid neutralizing
ability, lime (CaCOa) also appears to have the ability to reverse the deleterious
effects on soil structure that acidification can cause.
5.7.2 Soil Cations and Leaching
The suspected effects of acidification on the soil exchange complex have been
summarized by McFee (1982), Mortvedt (1982), and Tabatabai (1985). The effects docu-
mented in these references are as follows:
1. decrease in CEC as a result of clay alumination;
2. increase in the CEC of Ultisols as a result of sulphate adsorption;
3. decrease in base saturation and increase in soil acidity; and
4. increased formation of hydroxyl-Al interlayers under acid weathering.
None of these effects purported to be caused by acidic deposition have been demonstrated
convincingly (Turchenek et al . 1987). Lack of demonstrable cause-effect relationships
is particularly true for agricultural soils.
In a study of beech forest soils in Germany, Ulrich (1980) attributed a reduc-
tion in soil pH to wet and dry acidic (H^) deposition of approximately 1 kmol ha ^ y ^.
He also noticed a reduction in CEC which was ascribed to an increase in exchangeable Al .
Solubilized aluminum as well as exchangeable and Fe^^ also increased (Ulrich
1980). Because the study soil was already acidic (pH 3-4), any increase in acidity would
129
cause aluminum to solubilize in the form of either Al or A1(0H) cations. The perma-
nent negative charges on clay minerals found in this type of soil would cause the Al
polycations to be adsorbed even in preference to H"*" (Bohn et al . 1979). The adsorbed
aluminum and its cations in solution would be in equilibrium but would contribute to
soil acidity by means of hydrolysis.
Ross et al. (1985), in studies of a fertilized orchard in British Columbia,
found soil effects that mimicked acidic deposition. Their results indicated that
concerns regarding Ca and Mg nutrition and possible Al and Mn toxicity with respect to
fruit trees and fertilization are warranted. As noted previously, liming can counteract
these acidifying effects on soils, a mitigative procedure suggested by Ross et al.
(1985).
Although acidification to pH values lower than 3.5 is not likely under natural
conditions because of buffering by aluminum hydroxides and mineral weathering which
replenishes cations, simulated acidic rain experiments have demonstrated that aluminum
can be mobilized (Abrahamsen et al. 1976). Cronan and Schofield (1979) have also shown
that aluminum can be leached from forest soils into the aquatic environment under con-
ditions of acidic precipitation. However, it has been suggested by some investigators
that high concentrations of solubilized Al in soil and shallow groundwater may be natural
and related to climatic factors such as high rainfall rather than acidic deposition
(Krug and Frink 1983a, b; Nilsson and Bergkvist 1983).
The influence of acidic deposition on soil cation removal has been the subject
of a number of investigations. Some of these have been reviewed by Johnson et al.
(1983). Studies of the effects of acidic deposition on the following classes of soils
were reviewed: Inceptisols, Ultisols, and Spodosols in Washington, Tennessee, Alaska,
and Costa Rica. Base cation losses as a result of acidic atmospheric inputs were
considered to be insignificant in each of these studies (Johnson et al . 1983).
Natural cation losses as a result of leaching appear to be increased by:
organic acids in cold climates where the soils are undergoing podzolization; carbonic
acid in tropical and temperate soils; and nitric acid in nitrogen rich soils such as
those with nitrogen fixing vegetation (Turchenek et al. 1987).
A number of mathematical models formulated from similar concepts have been
developed to simulate chemical processes in soils that describe the influences of acidic
deposition on soil cations (Reuss 1980; Chri stopherson et al. 1982; Arp 1983; and
Gherini et al. 1985). Of these models, one developed by Reuss (1980) utilizes mass
balance equations in the simulation and, therefore, seems appropriate to discuss here.
His model predicts that a 1:1 removal of bases would occur in non-sulphate sorbing soils
under conditions of acidic deposition except in those soils with a very low base satura-
tion. In soils where base depletion was predicted, it was found that as base saturation
was depleted calcium removal decreased until the amount leached was in equilibrium with
atmospheric input. In arid areas where evapotranspi ration exceeds precipitation, the
model predicts that the ionic concentration of the soil solution would increase and that
the amounts lost via leaching would decrease. Calcium leaching was predicted by the
Reuss model to increase with increasing lime potential in soil systems with a high CO2
partial pressure. The model also predicted a dampening effect on cation loss in sulphate
adsorbing soils.
130
Three limitations of Reuss' (1980) model have been pointed out:
1. only calcium is considered in the model although other cations could be
included;
2. the release of cations to the exchange complex by weathering was not
considered;
3. the model is strictly abiotic and does not consider plant nutrient uptake.
5.7.3 Soil Anions
Anion sorption properties are important in terms of availability of plant
nutrients and for the regulation of the leaching rates of some elements. The general
order of affinity of soil for major anions is as follows:
H2P04~ > S04^~ > Cl" = NOa"
Nitrate and sulphate are of particular interest since they are both major constituent
ions of acidic deposition. Nitrate is not specifically adsorbed and its concentration
in the soil solution and its adsorption at a particular pH is governed by pH-charge
relationships; that is, little or no sorption occurs above the zero point of charge and
sorption increases with decreasing pH. Mott (1981) established that sulphate ions are
attracted by hydrous oxide surfaces more than nitrate or chloride ions. Sulphate
sorption has been demonstrated experimentally using sulphuric acid leaching, and it was
shown that iron podzols have a higher affinity for this ion than semi-podzol or brown
earth soils. Sulphate sorption has also been found to be high in iron and aluminum-rich
Inceptisols of Costa Rica while the Inceptisols, Utisols, and Spodosols of the United
States had a smaller affinity (Johnson et al. 1983).
Wiklander (1980) demonstrated that anions and soil type directly affect the
adsorption of cations in soil. For example, the Na, K, Ca, and Mg in association with
nitrate and chloride were bound less readily than when in association with phosphate in
soils. The same cations had an intermediate affinity for binding when in the presence
of sulphate (Wiklander 1980). Polyvalent anions added to soils increase adsorption and
decrease leaching of cations through this binding action. Additions of phosphate
fertilizers to soils can, therefore, reduce leaching losses of cations, especially if
the soils are rich in the hydrous oxides of aluminum and iron (Turchenek et al. 1987).
5.7.4 Availability of Nutrients and Toxic Metals
As part of their review, Turchenek et al . (1987) included a detailed documenta-
tion of nutrient cycling as it could be affected by acidic deposition. Since most of
their discussion deals with plant and microbial processes, the review synthesis is not
included here. The reader is, however, referred to Turchenek et al . (1987) should they
wish to cross-link findings from all three major fields of research. Other major
reviews of this topic may be found in Mortvedt (1982) and Tabatabai (1985).
131
5.8 EFFECTS OF ANTHROPOGENIC SOURCES OF ACIDITY
5.8.1 Nitrogenous Fertilizers
The gradual acidification of soils as a result of the nitrification of ammonium
based fertilizers has been extensively studied (citations in McCoy and Webster 1977). A
review of the reaction chemistry of nitrogenous fertilizers with soils and their compon-
ents may be found in Penney and Henry (1976). Urea application causes an initial
increase in soil pH by forming NHa which is then oxidized by the bacterium Nitrosomonas
to form NH4^ and nitrite (NO2 ). The oxidation process (NHa to NH4^) produces 2 moles H^
per mole of NH4''". Nitrite is then converted by Nitrobacter to nitrate (NOa").
Ammonium sulphate has the highest equivalent acidity of the various nitrogen
based fertilizers and requires a CaCOatN ratio of 5.35 for neutralization (Penney and
Henry 1976). In comparison, anhydrous ammonia, urea, and arranonium nitrate require a
ratio of only 1.80 for neutralization. The actual impact of the acidifying process is
determined by a variety of factors such as proportion of fertilizer taken up by plants
prior to nitrification, NH4'*' and NOa~ concentration, pH of the soil, oxygen supply, and
temperature (Alexander 1977). Under field conditions, the fixation and leaching of
nitrogenous ions and the influences of these ions and plant growth on base mobilization
determine the amount of acidity actually produced. Because of the complexity of the
above processes and their interdependence, acidification as a result of nitrogen
fertilizer application under field conditions has usually been determined empirically.
McCoy (1973) and Penney and Henry (1976) have reviewed this topic.
The problem of acidity in agricultural soils of the Western provinces of Canada
and in particular. Alberta, has been reviewed by Penney et al. (1977) and Hoyt et al.
(1981). The primary cause of acidification in most instances was fertilizer usage and
not industrial emission-caused acidic deposition, although this was cited as another
possible cause. Ross et al. (1985) reported the acidification of orchard soils in the
Okanagan Valley of British Columbia as a direct result of fertilization. Numerous other
studies (McCoy and Webster 1977; Hoyt et al. 1981; and Nyborg and Mahli 1981) have
substantiated the effects of fertilization on soil acidity, which include lowering of
surface and subsurface pH, aluminum solubilization, and consequent leaching.
Liming of soils seems to prevent the acidification process even when fertiliza-
tion practices continue. It has been suggested that liming should become a general
practice to combat soil acidification (Hoyt et al. 1981). These authors also concluded
that, if in general liming is practiced to fight fertilizer caused acidity in soils, it
would negate any effects of the deposition of acid forming compounds. Specifically with
regard to agricultural soils, Turchenek et al. (1987) felt that SO2 and NOx would not
threaten agricultural productivity if liming practices became general and, in fact, the
addition of these plant nutrients via atmospheric deposition may be beneficial. A study
conducted by Sauerbeck (1983) in West Germany tends to support the liming hypothesis.
His results showed that local S deposition rates at his study sites varied from 20 to
100 kg S ha ^ but, as a result of liming, soil acidification did not occur and little
damage to field crops resulted. Sauerbeck also found that liming caused cations that
normally would have leached from the soils under acidifying conditions to be retained.
132
5.8.2 Atmospheric Deposition
The main acid forming atmospheric pollutants in Alberta are anthropogenic
emissions of SOx, reduced sulphur compounds, and NOx (Hunt et al. 1982). The removal
from the atmosphere of these compounds falls into two general categories - wet and dry
deposition. The various mechanisms for each type of deposition have been categorized
by Fowler (1980), Galloway and Parker (1980), and Krupa et al. (1987), and are as
f ol lows :
1 . Wet Deposition
a. Incident wet deposition - gravitational transfer of water to earth
surfaces ;
b. Throughfall - water that has passed through a leaf canopy;
c. Net throughfall - difference between throughfall and incident depo-
sition for a particular area.
2. Dry Deposition
a. Dry fall - soil and seasalt particles of, for example, >10-30 pm
diameter which settle by gravity;
b. Aerosol impaction - particles of <3 ym diameter which deposit on to
surfaces. These are commonly particles of (NH4)2S04, NH4NO3, and
others ;
c. Gaseous adsorption - gases sorbed by foliage or soils (i.e., SO2,
NOx, and CO2).
Recent reviews of deposition mechanisms and interactions of gaseous and
particulate materials in relation to their effects on plant and soil surfaces have been
provided by Fowler (1980), Chamberlain (1986), and Weidensaul and McClenahen (1986).
The resistance concept as reported in Fowler (1980) has been used to mathemati-
cally describe the flux of atmospheric deposition to surfaces such as soils. The actual
surface resistances of soils have not been measured, however, due to their extremely
heterogenous nature. It is known that resistance increases with decreasing soil pH and
increasing dryness. More gases will be sorbed by soils with high moisture content or
high calcium content since both properties exhibit almost negligible resistance factors.
This is true for the major gases and/or* vapours SO2, NO2, and HNOa" (Turchenek
et al. 1987). Other gases that can be adsorbed during the deposition process include
nitrous oxide, nitric oxide, hydrogen sulphide, methyl mercaptan, dimethyl disulphide,
carbonyl sulphide, and carbon disulphide (Committee on the Atmosphere and the Biosphere,
U.S. 1981).
Gaseous vapours and particulate materials can be incorporated into cloud
droplets with deposition occurring as a precipitation event. The adsorption and
oxidation of acid forming gases by cloud and rain droplets has been described by Fowler
(1980) and Krupa et al. (1987).
Very little in the way of hard data is available for Alberta with respect to
dry deposition. Caiazza et al . (1978) found dryrwet deposition ratios of 4.8 for
sulphate and 2.1 for total nitrogen in the Edmonton area. Dry deposition inputs to a
mixed hardwood forest in the eastern United States have been found to be much more
important than had previously been thought (Lindberg et al. 1986). While wet deposition
133
was the primary source for S04^~ and NHa^, dry deposition was most important for fine
particle or vapour NOa" and and for coarse particle and Ca^ .
Generally, soils have a high sulphur sorption capacity and the reactions of
atmospherically deposited sulphur with soils have been reviewed extensively by Chaudry
et al. (1982) and by others. Sulphur sorption is influenced to a major extent by soil
moisture content. Nyborg et al. (1977, 1980) and Hsu and Hodgson (1977) have demon-
strated experimentally that Alberta soils have a high capacity for sulphur sorbtion and
that this can result in pH depression. In their studies, Nyborg et al . (1977) found
increases in sulphur of 12 to 53 kg ha ^ downwind of SO2 sources. The highest
sulphur gains in these experiments were at sites close to, or farthest from, the source
at a distance of 2 and 37 km. Generally, little of the sorbed sulphur was in the form
of sulphate. Evaluation of the other potential sources such as snow pack melt, rainfall,
and surface waters indicated that the primary source of sulphur was dry and not wet
deposition, although the latter contributed to the overall loading.
The results of the aforementioned studies of SO2 deposition and acidification
were only approximations but do indicate a potential for the acidification of Alberta
soils. Some limitations of these studies were pointed out by Turchenek et al. (1987)
and are as follows:
1. the method of total S determination in the study was imprecise;
2. pH depressions reported of 0.1 and 0.2 units were within the temporal and
normal spatial variability for the type of soil studied;
3. extrapolation errors involved in converting small sample plot data to a
kg~^ ha ^ loading may have been large;
4. the S forms involved were not identified.
In most natural and agricultural ecosystems, air pollutants will first encounter
plant canopies. Plant canopies have a tremendous SO2 sorption capacity which diminishes
with the time of exposure. This topic has been discussed at some length in a previous
major section of this report. However, in any study of the effects of acidic deposition
on soils, vegetation processes must be taken into account in an integrative way for
interpreting the data or for precise determinations of impacts. This point was also
made by the study group of Turchenek et al. (1987) whose document forms the basis of the
soils portion of this report.
5.9 SUMMARY
A summary of the potential impacts of acidic deposition on soils is given in
Table 30 which was derived from Turchenek et al. (1987), McFee (1982), and Cook (1983).
With respect to agricultural lands, Table 31 shows the estimated amounts of cultivated
land in different ranges of soil pH for the Canadian Great Plains (Turchenek et al.
1987).
134
Table 30. Summary of the potential impact of acidic deposition on
soi Is.
Process or Property
Hypothetical Impact of Acid Deposition
I. Soil Exchange Complex
Exchange Capacity
Exchangeable Acidity
Base Saturation
Clay Mineral
Morphology
Aluminum
II. Organic Matter
Organic Matter
Turnover
Microbial Community
Dynamics
Organo-Mineral
Associations
Root Uptake
III. Plant Nutrients
Nitrogen
Decrease in CEC resulting from clay
alumination
Increase in CEC of soils with oxy-
hydroxides due to sulphate
adsorption
Increase
Decrease
Increased formation of hydroxy-Al
interlayers and acid weathering
Increased mobilization and leaching
Increased availability and toxicity
Decreased rate of C mineralization due
to acidification and/or associated
trace metal toxicity
Decreased CO2 flux from land to
atmosphere
Increased retention of organic matter
Shift from bacteria to more acid -
tolerant fungi
Reduced organo-clay interaction due to
disruption of cation bridge linkages
Trace metal toxicity due to
acidification
Decreased ammonif ication
Decreased nitrification
Changes in products of denitrif ication
Increase in leaching
Enhanced cation leaching due to NOa"
inputs
Reduced plant availability
continued.
135
Table 30. (Concluded).
Process or Property
Hypothetical Impact of Acid Deposition
Sulphur
Increased S04^" reduction in low S,
anoxic systems
Increased reduced-S flux and reduced
CH4 flux to atmosphere
Decreased leaching of S
Decreased leaching of cations in
sesquioxidic soils; increased
leaching in others
Reduced plant availability
Phosphorus
Decreased leaching and A1P04
precipitation in soil with high Al
Increased P043~ solubilization, plant
availability and leaching in
calcareous soils
Reduced availability with pH reduction
Fe, Mn, Zn, Cu, Co
Increased availability
Increased leaching
Mo, B
Reduced availability
Ca, Mg, K
Reduced availability
Increased leaching
Toxic Elements
Some micronutrients may reach toxic
levels due to increased solubility
Increased concentrations, toxicity
and leaching of heavy metals
Increased Al toxicity
IV Weathering
Carbonates
Increased dissolution
Primary Minerals
Increased dissolution
Clay Minerals
Increased alumination (formation of
Al interlayers)
Reduced surface charge
136
Table 31. Estimated amounts of cultivated land in different
ranges of soil pH on the Great Plains.^
Province or Region Hectares (x 1000)
pH <5.5 pH 5.6-6.0 pH 6.1-6.5
Manitoba 10*
Saskatchewan 202 202 405
Alberta, excluding the
Peace River region^ 230 1166 2276
Peace River region
of Alberta and 98 447 602
British Columbia^
TOTAL 530 1825 3283
^Adapted from the original table in Hoyt et al. (1981).
2pH <6.0.
^Based on soil samples taken during 1962-72.
137
5.10 EFFECTS OF ACIDIC DEPOSITION ON SOILS: LITERATURE CITED
Abrahamsen, G., J. Hovland, and S. Hagvar. 1980. Effects of artificial acid rain and
liming on soil organisms and the decomposition of organic matter. In: Effects
of Acid Precipitation on Terrestrial Ecosystems. NATO Conference Series.
Volume 4, eds. T.C. Hutchinson and M. Havas. New York: Plenum Press,
pp. 341-362.
Abrahamsen, G., K. Bjor, R. Horntvedt, and B. Tveite. 1976. Effects of acid precipitation
on coniferous forest. SNSF Research Report 6: 37-63.
Alexander, M. 1977. Introduction to Soil Microbiology, 2nd ed. New York: John Wiley and
Sons. 467 pp.
Arp, P. A. 1983. Modelling the effects of acid precipitation on soil leachates: a simple
approach. Ecological Modelling 19: 105-117.
Bache, B.W. 1980. The acidification of soils. In: Effects of Acid Precipitation on
Terrestrial Ecosystems. NATO Conference Series. Volume 4, eds. T.C. Hutchinson
and M. Havas. New York: Plenum Press, pp. 183-202.
Bache, B.W. 1979a. Soil reaction. In: The Encyclopedia of Soil Science Part I. Physics,
Chemistry, Biology, Fertility, and Technology, eds. R.W. Fairbridge, and
C.W. Finkl. Stroudsburg, Pennsylvania: Dowden, Hutchinson and Ross,
pp. 487-492.
Bache, B.W. 1979b. Base saturation. In.: The Encyclopedia of Soil Science Part I. Physics,
Chemistry, Biology, Fertility, and Technology, eds. R.W. Fairbridge and
C.W. Finkl. Stroudsburg, Pennsylvania: Dowden, Hutchinson and Ross, pp. 38-42.
Bache, B.W. 1976. The measurement of cation exchange capacity of soils. Journal of the
Science of Food and Agriculture 27: 273-280.
Barratt, B.C. 1970. Effect of long-term fertilizer topdressings in hay fields on humus
forms and their micromorphology . Agri Digest 21: 11-18.
Black, C.A. 1968. Soil-Plant Relationships. New York: John Wiley and Sons. 792 pp.
Bohn, H., B. McNeal, and G. O'Connor. 1979. Soil Chemistry. New York: John Wiley and
Sons. 329 pp.
Buol, S.W., F.D. Hole, and R.J. McCracken. 1980. Soil Genesis and Classification, 2nd
ed. Ames: Iowa State University Press. 404 pp.
Burchill, S., D.J. Greenland, and M.H.B. Hayes. 1981. Adsorption of organic molecules.
In: The Chemistry of Soil Processes, eds. D.J. Greenland and M.H.B. Hayes.
Chichester: John Wiley and Sons, pp. 221-400.
Caiazza, R., K.D. Hage, and D. Gallup. 1978. Wet and dry deposition of nutrients in
Central Alberta. Water, Air, and Soil Pollution 9: 309-314.
Chamberlain, A.C. 1986. Deposition of gases and particles on vegetation and soils.
In: Air Pollutants and Their Effects on the Terrestrial Ecosystem, eds.
A.H. Legge and S.V. Krupa. New York: John Wiley and Sons, pp. 189-209.
Chaudry, M., M. Nyborg, M. Mol ina-Ayala , and R.W. Parker. 1982. The reaction of SO2
emissions with soils. I_n: Acid Forming Emissions in Alberta and Their Ecological
Effects. Symposium Workshop Proceedings. Alberta Environment, Canadian
Petroleum Association, and the Oil Sands Environmental Study Group. 1982 March
9-12; Edmonton, Alberta; pp. 415-433.
Christophersen, N., A.O. Stuanes, and R.F. Wright. 1982. Runoff chemistry at a mini-
catchment watered with "unpolluted precipitation." Nordic Hydrology 13:
115-128.
Committee on the Atmosphere and the Biosphere. 1981. Atmosphere-Biosphere Interactions:
Toward a Better Understanding of the Ecological Consequences of Fossil Fuel
Combustion. Washington, DC: National Academy Press. 263 pp.
138
Cook, R.B. 1983. The impact of acid deposition on the cycles of C, N, P, and S. rn: The
Major Biogeochemica! Cycles and Their Interactions, eds. B. Bolin and
R.B. Cook. Chichester: John Wiley and Sons, pp. 345-364.
Cronan, C.S. and C.L. Schofield. 1979. Aluminum leaching response to acid precipitation:
Effects on high elevation watersheds in the northeast. Science 204: 304-306.
Finkl, C.W. 1979. Leaching. In: The Encyclopedia of Soil Science Part I. Physics,
Chemistry, Biology, Fertility, and Technology, eds. R.W. Fairbridge and C.W.
Finkl. Stroudsburg, Pennsylvania: Dowden, Hutchinson and Ross, pp. 260-265.
Fowler, D. 1980. Wet and dry deposition of sulphur and nitrogen compounds from the atmos-
phere. In.: Effects of Acid Precipitation on Terrestrial Ecosystems. NATO
Conference Series. Volume 4, eds. T.C. Hutchinson and M. Havas. New York:
Plenum Press, pp. 9-27.
Galloway, J. N. and G.G. Parker. 1980. Difficulties in measuring wet and dry deposition
on forest canopies and soil surfaces. I_n: Effects of Acid Precipitation on
Terrestrial Ecosystems. NATO Conference Series Volume 4, eds. T.C. Hutchinson
and M. Havas. New York: Plenum Press, pp. 57-68.
Gessa, C. 1979. Buffers, Buffering. In: The Encyclopedia of Soil Science Part I. Physics,
Chemistry, Biology, Fertility, and Technology, eds. R.W. Fairbridge and
C.W. Finkl. Stroudsburg, Pennsylvania: Dowden, Hutchinson and Ross, pp. 42-47.
Gherini, S.A., Lingfung Mok, R.J.M. Hudson, G.F. Davis, C.W. Chen, and R.A. Goldstein.
1985. The ILWAS model: formulation and application. Water, Air and Soil
Pollution 26: 425-459.
Goh, T.B., M.J. Dudas, S. Pawluk, and J.T. Harapiak. 1986. Micromorphological , physical
and chemical effects of nitrogen fertilizers on soils. Canadian Society of
Soil Science, Titles and Abstracts for the Annual Meeting. 1986 July 7-10;
University of Saskatchewan, Saskatoon; p. 31.
Hamblin, A. P. and D.J. Greenland. 1977. Effect of organic constituents and complexed
metal ions on aggregate stability of some East Anglian soils. Journal of Soil
Science 28: 410-416.
Helling, C.S., G. Chesters, and R.B. Corey. 1964. Contribution of organic matter and
clay to soil cation exchange capacity as affected by the pH of the saturating
solutions. Soil Science Society of America Proceedings 28: 517-520.
Holowaychuk, N. and J.D. Lindsay. 1982. Distribution and relative sensitivity to acidi-
fication of soils, Sand River area, Alberta. Prep, for the Research Management
Division, Alberta Environment and Canadian Petroleum Association. RMD Report
82/13. Edmonton, Alberta. 93 pp.
Hoyt, P.B., M. Nyborg, and H. Ukrainetz. 1981. Degradation by acidification. Proceed-
ings of the 18th Annual Soil Science Workshop. 1981 February 24-25; Edmonton,
Alberta; pp. 41-71.
Hsu, H. and G.W. Hodgson. 1977. Organic compounds of sulphur: Initial data for soils
and streams in Alberta. Ijn: Proceedings of Alberta Sulphur Gas Research
Workshop III, eds. H.S. Sandhu and M. Nyborg. 1977 November 17-18; University
of Alberta, Edmonton. Edmonton, Alberta: Research Secretariat, Alberta
Environment; pp. 246-263.
Hunt, J.E., R.G. Wright, and R.L. Desjardins. 1982. A measurement system for SO2 dry
deposition rates. In: Pre-Proceedings . Acid Forming Emissions in Alberta and
Their Ecological Effects. 1982 March 9-12; Edmonton, Alberta; pp. 57-88.
Jackson, M.L. 1958. Soil Chemical Analysis. Englewood Cliffs, N.J.: Prentice-Hall.
198 pp.
Jenkinson, D.S. 1981. The fate of plant and animal residues in soil. In: The Chemistry
of Soil Processes, eds. D.J. Greenland and M.H.B. Hayes. Chichester: John
Wiley and Sons, pp. 505-561.
139
Jenny, H. 1941. Factors of Soil Formation. New York: McGraw-Hill Book Co. 281 pp.
Johnson, D.W., D.D. Richter, H. Van Miegroet, and D.W. Cole. 1983. Contributions of acid
deposition and natural processes to cation leaching from forest soils: A review.
Journal of the Air Pollution Control Association 33: 1038-1041.
Kalisz, P.J. and E.L. Stone. 1980. Cation exchange capacity of acid forest humus layers.
Soil Science Society of America Journal 44: 407-413.
Krug, E.C. and C.R. Frink. 1983a. Effects of acid rain on soil and water. Bulletin 811.
New Haven: The Connecticut Agricultural Experiment Station. 45 pp.
Krug, E.C. and C.R. Frink. 1983b. Acid rain on acid soil: A new perspective. Science
221: 520-525.
Krupa, S.V., V.C. Bowersox, B.I. Chevone, W.W. Knapp, J. A. Lynch, and W.W. McFee. 1987.
Analysis of Inter-regional Project IR-7 Precipitation Chemistry Data. Precipi-
tation Chemistry in the United States. Summary of Qualitative Characteristics.
1979-1984. Progress Report Submitted to USDA-Cooperati ve State Research
Service, Washington, DC. 211 pp.
Lindberg, S.E, G.M. Lovett, D.D. Richter, and D.W. Johnson. 1986. Atmospheric deposition
and canopy interactions of major ions in a forest. Science 231: 141-145.
McCoy, D.A. 1973. Some effects of long-term fertilizer application on the Breton plots.
Edmonton: University of Alberta. 116 pp. M.Sc. Thesis.
McCoy, D.A. and 6.R. Webster. 1977. Acidification of a Luvisolic soil caused by low-rate,
long-term applications of fertilizers and its effects on growth of alfalfa.
Canadian Journal of Soil Science 57: 119-127.
McFee, W.W. 1982. Sensitivity ratings of soils to acid deposition: A review. I_n: Response
of Agricultural Soils to Acid Deposition, Workshop Proceedings, eds, M.F. Arthur
and C.K. Wagner. Columbus: Battelle Columbus Laboratories: pp. 2-1 - 2-12.
Mortvedt, J.J. 1982. Impacts of acid deposition on micronutrient cycling in agroecosys-
tems. In: Response of Agricultural Soils to Acid Deposition, Workshop
Proceedings, eds. M.F. Arthur and C.K. Wagner. Columbus: Battelle Columbus
Laboratories, pp. 2-64 - 2-75.
Mott, C.J.B. 1981. Anion and ligand exchange. In: The Chemistry of Soil Processes, eds.
D.J. Greenland and M.H.B. Hayes. Chichester: John Wiley and Sons, pp. 179-219.
Nilsson, S.I. and B. Bergkvist. 1983. Aluminum chemistry and acidification processes in
a shallow Podzol on the Swedish westcoast. Water, Air, and Soil Pollution 20:
311-329.
Nyborg, M. and S.S. Mahli. 1981. Methods of reducing the soil acidification caused by
nitrogen fertilizers. Xn: Proceedings of the Saskatchewan Advisory Council and
Crops Workshop; Saskatoon: University of Saskatchewan; pp. 112-118.
Nyborg, M., R.W. Parker, L.W. Hodgins, D.H. Laverty, and S. Tayki . 1980. Soil acidifica-
tion by SO2 emissions in Alberta, Canada. In: Ecological Impact of Acid
Precipitation, Proceedings of an International Conference, eds. D. Drablos and
A. Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway;
pp. 180-181.
Nyborg, M., J. Crepin, D. Hocking, and J. Baker. 1977. Effect of sulphur dioxide on pre-
cipitation and on the sulphur content and acidity of soils in Alberta, Canada.
Water, Air, and Soil Pollution 7: 439-448.
Oades, J.M. 1984. Soil organic matter and structural stability: mechanisms and implica-
tions for management. Plant and Soil 76: 319-337.
140
Oades, J.M. and M.N. Ladd. 1977. Biochemical properties: carbon and nitrogen metabolism.
In: Soil Factors in Crop Production in a Semi-Arid Environment, eds.
J.S. Russell and E.L. Greacen. St. Lucia: University of Queensland Press, in
association with the Australian Society of Soil Science, pp. 127-160.
Paul, E.A. 1970. Plant components and soil organic matter. Recent Advances in Phyto-
chemistry 3: 59-104.
Peech, M. 1965. Hydrogen ion activity. In: Methods of Soil Analysis, Part 2, eds. C.A.
Black, D.D. Evans, J.L. White, L.E. Ensminger, and F.E. Clark. Madison,
Wisconsin: American Society of Agronomy, pp. 914-932.
Penney, D.C. and J.L. Henry. 1976. Reaction of fertilizer N with soil components. Pro-
ceedings of the 14th Annual Soil Science Workshop. 1976 January 20-21;
Calgary, Alberta; pp. 275-298.
Penney, D.C, M. Nyborg, P.B. Yout, W.A. Rice, B. Siemens, and D.H. Laverty. 1 977. An
assessment of the soil acidity problem in Alberta and northeastern British
Columbia. Proceedings of the 15th Annual Soil Science Workshop. 1977 February
1-2; Edmonton, Alberta; pp. 179-189.
Reuss, J.O. 1985. Keynote address: Modeling the effect of acid deposition on soil and
water acidification. In: Predicting Soil and Water Acidification. Proceedings
of a Workshop, Knoxville, Tennesse, ed. D.W. Johnson. 1984 March 26-29; Oak
Ridge, Tennessee; pp. 3-14.
Reuss, J.O. 1980. Simulation of soil nutrient losses resulting from rainfall acidity.
Ecological Modelling 11: 15-38.
Ross, 6.J., P.B. Hoyt, and G.H. Neilson. 1985. Soil chemical and mineralogical changes
due to acidification in Okanagan apple orchards. Canadian Journal of Soil
Science 65: 347-355.
Sauerbeck, D. 1983. Auswirkung des, "sauren Regens" auf landwi rtschaf tl ich genutzte
Boden. Landbauf orschung Volkenrode 33: 201-207.
Schnitzer, M. 1980. Effect of low pH on the chemical structure and reactions of humic
substances. In: Effects of Acid Precipitation on Terrestrial Ecosystems.
NATO Conference Series Volume 4, eds. T.C. Hutchinson and M. Havas. New York:
Plenum Press, pp. 203-222.
Schnitzer, M. 1978. Humic Substances: Chemistry and Reactions. In: Soil Organic Matter,
eds. M. Schnitzer and S.U. Khan. Amsterdam: Elsevier Scientific Publishing,
pp. 1-64.
Schofield, R.K. and A.W. Taylor. 1955. The measurement of soil pH. Soil Science Society
of America Proceedings 19: 164-167.
Stevenson, F.J. 1982. Humus Chemistry Genesis, Composition, Reactions. New York: John
Wiley and Sons, 443 pp.
Tabatabai , M.A. 1985. Effect of acid rain on soils. Critical Reviews in Environmental
Control 15: 65-110.
Talibuden, 0. 1981. Precipitation. Xn: The Chemistry of Soil Processes, eds. D.J.
Greenland and M.H.B. Hayes. Chichester: John Wiley and Sons, pp. 115-177.
Thomas, G.W. and W.L. Hargrove. 1984. The chemistry of soil acidity. In: Soil Acidity
and Liming, ed. F. Adams. Madison: American Society of Agronomy, Inc., Crop
Science Society of America, Inc., and Soil Science Society of America, Inc.,
pp. 3-56.
Tisdall, J.M. and J.M. Oades. 1982. Organic matter and water-stable aggregates in soils.
Journal of Soil Science 33: 141-163.
141
Turchenek, L.W., S.A. Abboud, C.J. Tomas, R.J. Fessenden, and N. Holowaychuk . 1987.
Effects of Acid Deposition on Soils in Alberta. Prep for the Acid Deposition
Research Program by the Alberta Research Council, Edmonton. ADRP-B-05-87 .
202 pp.
Ulrich, B. 1980. Production and consumption of hydrogen ions in the ecosphere. In:
Effects of Acid Precipitation on Terrestrial Ecosystems. NATO Conference
Series Volume 4, eds. T.C. Hutchinson and M. Havas, New York: Plenum Press,
pp. 255-282.
van Breemen, N., C.T. Driscoll, and J. Mulder. 1984. Acidic deposition and internal
proton sources in acidification of soils and waters. Nature 307: 599-604.
van Breemen, N., J. Mulder and C.T. Driscoll. 1983. Acidification and alkalinization
of soil. Plant and Soil 75: 238-308.
Weidensaul, T.C. and J.R. McClenahen. 1986. Soil-air pollutant interactions. In: Air
Pollutants and Their Effects on the Terrestrial Ecosystem, eds. A.H. Legge and
S.V. Krupa. New York: John Wiley and Sons, pp. 397-414.
Wiklander, L. 1980. Interactions between cations and anions influencing adsorption and
leaching. In: Effects of Acid Precipitation on Terrestrial Ecosystems. NATO
Conference Series, Volume 4, eds. T.C. Hutchinson and M. Havas. New York:
Plenum Press, pp. 239-254.
Wiklander, L. 1979. Leaching and acidification of soils. In: Ecological Effects of Acid
Precipitation. Report of Workshop 6atehouse-of-Fleet , Galloway, UK, EPRI
EA-79-6-LD. SOA 77-403, ed . M.J. Wood. 1978 September 4-7; Palo Alto, CA:
Electric Power Research Institute, pp. 1-21.
142
143
6. EFFECTS OF ACIDIC DEPOSITION ON SOIL MICROORGANISMS AND MICROBIALLY MEDIATED
PROCESSES.
6.1 INTRODUCTION
The soil microbial community is comprised of six main groups: actinomycetes ,
algae, bacteria, fungi, protozoa, and micro-invertebrates. The microbial biomass is
dominated by the bacteria and fungi, forming 25% and 75% of the total biomass, respec-
tively (Anderson and Domsch 1978). It has been estimated that up to 90% of the
respiratory metabolism of the soil-litter system is due to bacterial and fungal activity
(Reichle 1977; Persson et al. 1980). For these reasons, much of the research regarding
acidic impacts on soil biology has been directed to the bacteria and fungi and their
functions .
The major roles of the soil microbial community in maintaining soil fertility
have been summarized by Alexander (1980):
1. The transformation of soil nitrogen, phosphorous, and sulphur from organic
to inorganic forms, resulting in the availability of these nutrients for
utilization by plants;
2. The formation of humus which improves soil structure and promotes root
growth due to better aeration and moisture conditions. Soil structure is
also improved by the creation of soil aggregates which result from the
binding properties of microbial excretions and fungal mycelial production;
3. The decomposition of organic matter and the detoxification of phytotoxins
resulting from anaerobic decay processes; and
4. The rapid degradation of toxic substances introduced into the soil includ-
ing pesticides, herbicides, sewage sludge, and carbon monoxide.
In addition, specialized bacteria and fungi which form symbiotic associations
with the roots of many of the higher plants are vital in improving plant nutrition.
Examples of such symbiotic organisms are nitrogen-fixing bacteria which form root
nodules (Rhizobium) and mycorrhizal fungi which are associated with the roots of 95% of
the higher plants and which serve in mobilizing plant nutrients not easily available to
root systems.
The effects of acidic deposition on soil microorganisms were speculated to be
either acute, resulting from high dosages over short periods, or chronic, resulting from
low dosages over long periods of time, i.e., years with periodic, intermittent episodes
(Visser et al. 1987).
6.2 GENERAL EFFECTS OF ACIDIC DEPOSITION ON SOIL MICROBES
Acidic deposition may have both direct and indirect effects on soil microbes.
Direct effects include:
1. An increase in the hydrogen ion concentration to levels where cell membrane
permeability and enzyme systems located at the cell surface are altered.
At relatively low pH levels, undi ssociated acids can enter the cells and
affect them by changing the internal cell pH; and
144
2. The production of soluble anions may or may not be toxic to microorganisms.
For example, Babich and Stotzky (1978) tested a number of fungi and
bacteria for their sensitivity to bisulphite and sulphite, and observed
that the HSOa was more toxic, particularly at pH values below 6.0
where this anion dominated. They argued that since all microbial cells
are covered with a thin film of water, the toxic effects of sulphur dioxide
could be due to its solubility at high pH and the formation of bisulphite
and sulphite on the cell surfaces.
Indirect effects include:
1. An increase in leaching of nutrient cations due to their replacement by
hydrogen ions at exchange sites. The resultant nutrient deficiencies may
alter microbial growth and reproduction;
2. Increase in solubility of aluminum and iron to levels where they become
toxic to microorganisms;
3. A modification in the physiology of plants to a degree where root symbionts
become less effective. In addition, changes in chemical and physical
properties of the soil as mentioned previously, may adversely influence
plant growth with consequential effects on the root symbionts; and
4. A modification of the nutrient quality of plant residues available to the
decomposer community. Uptake of sulphur dioxide by the plant, dry
deposition of sulphur dioxide on plant surfaces, and enhanced uptake of
nutrients such as aluminum and iron due to altered soil chemistry may
change the chemical quality of plant residues which may indirectly
influence the microbial community.
6.2.1 Influence of Soil Acidity on Microbial Communities
The causes of natural soil acidity have been fully discussed in other sections of this
review. The effects of anthropogenic changes to the natural soil acidity have also been
reported. In general, the effects of lowered soil pH on the microbial community can be
summarized as follows:
1. Soil bacteria, except for S-oxidizing bacteria (Thiobaci 1 lus thioxidans) ,
and actinomycetes are inhibited below pH 5.0;
2. Fungi show no sensitivity to pH change unless the resultant pH is extremely
high or low. In fact, at pH values that inhibit bacterial activity, fungi
flourish because of a lack of competition. Protozoans appear to have no
marked sensitivity to pH except under conditions that also affect fungi;
3. Blue-green algae are sensitive to pH values below 5.0 while the green
algae appear to be less sensitive;
145
4. All microbially mediated processes, with the exception of S-oxidation,
operate optimally at neutral or nearly neutral pH values. Specialized
bacteria performing nitrogen fixation and nitrification appear to be
sensitive to acidity while fungal processes such as ammonif ication and
decay are not (bacteria performing such processes are, however, sensi-
tive); and
5. Sulphur oxidation appears to be insensitive to soil pH since it can be
performed by acid tolerant bacteria and fungi.
A summary of the general effects of soil pH on microbial processes is presented
in Table 32.
6.2.2 Summary of Acidic Deposition Effects on Microbial Processes
In their review of the effects of acidic deposition on soil microorganisms,
Visser et al. (1987) reviewed a major portion of the current literature on the topic.
They concluded that it is difficult to determine if soil acidification caused by acidic
deposition has effects on the soil microflora, or the processes they mediate. Most of
the studies reviewed by Visser et al. were conducted using simulated acidic rain and
showed widely varying effects on the various microbial components of the soil community.
The levels at which changes in soil pH begin to affect the various microbial types were
summarized as follows:
Nitrogen fixation
Nitri f ication
Ectomycorrhi zae
Vesicular-arbuscular mycorrhizae
Organic decomposition
Ammonif ication
Soil respiration
Carbon mineralization
Community structure
Soil enzymes
pH 6.0
pH 6.0
pH not certain
pH 6.0
pH 2-4
pH 3.0
pH 3.0
pH 3.0
pH 3-4
pH not certain but possibly 2.0
The majority of studies reviewed by Visser et al . (1987) were conducted on
naturally acidic forest soils where microflora adapted to acidity exist. Therefore,
these results may be somewhat misleading if one is considering grassland and
agricultural systems.
Based on the results of both laboratory and field studies reviewed by Visser et
al. (1987), it appears that a pH reduction of naturally acidic forest soils to 3.0 or
less would inhibit soil respiration. Due to the high buffering capacity of decaying
plant residues and organic matter in the forest floor and the probable adjustment of the
microflora previously adapted to acidity to further acidification, acidic rain of at
least pH 2.0 or high dosage rates of sulphur dioxide or elemental sulphur would be
necessary to reduce the soil pH to an extent sufficient to alter microbial respiration.
It is unclear whether or not this would also be the case in agricultural systems.
146
a: fo
O)
>>
Q. -M
ro
u
<— >,+->
c:
•r—
to
E (O x:
>>
ro
ro
>
a>
^ i- O.XZ
•r-
J-
•
•p-
o
O)
E +J o
O.
•
l.
QJ
+-»
>
-H
•r- 3 i-
O
>>
Q>
1 —
OJ
•p-
E
a>
+-> <D O
J-
i~
•
+->
O
SZ
to
o
o
Q. C 1—
o
O
>>
-a
O
+->
o>
C
■H
>>
o x:
r—
+->
■H
• c
ro
ez
•r-
QJ
Q
E
O C_5
SZ
•r-
•f—
to ro
X3
•1—
SZ
to
O)
o
c +-»
<_)
XI
SZ
XT
+->
ro
L.
c
to
•1 —
3
Q. 1
ro E
QJ S-
i_
□c
-M
•f—
-o •
OJ
x:
E
O 01
O
+-> O
o
Q.
(/)
4-»
01 0) to
1.
c
E
3 i-
ro SZ
O
C +-> (O
o
•f—
o
+-> O
O 4-
i- a
3:
• •>
to
•r- •!- <U
o
O (/)
+-> o
Q.
ac
> E ^
to
■M
(/) -f-
tn
Q.
to
to '1— to
c
p—
3 O
•
o
c: TD
■H
SZ t—
o
ro
ro -H
in
O
o o
ro
CO •
3
c
to
•r—
O
•
-H +->
E E
l_
>>
o
Q)
CI <U 3
+-»
+-»
L-
E 00
CO
•r—
Q) to
+->
nr +->
1—
E
•r- fO O
•1—
(U
0)
>> +->
■O ro
3
Q.«i-
c:
to KJi <D
1
■a
+-»
x: 1 —
1
+J O)
t—
QJ
sz
M-
o •
i~ 1— i-
U
a ro
•P- O.
>) Q-
C
•H T-
o
i- +->
■H <a «o
to
o
ro
+->
o
o E
ro
rO p-
•r- J/>
to O
f —
o
X)
1 c:
ro O
E •
+->
ro
•M
> 't-
C r—
0)
CM
o. u
CD ro
QJ ^
C
C X
r— Q) <0
>
u
M-
ro E
ro
to
o p-
ro
a> 0)
(0 0)0
•p-
o
•1- QJ
3C
o u-
QJ QJ
+-»
3 ro
i-
•~
■o
i- 1 —
O.
O
to
■o
O)
o o
■a cn c
•f—
(U
0) QJ
QJ
<_) to
QJ
o i-
r —
•r- "O
•1- 1 T-
o
Nl
+->
c
> ^
Q) C
X)
S- o
o
•a
> O)
cx
ro
•f—
U CT
QJ
ro <_)
Q. QJ
o.
-»->
•1— to
•1- 3 to
to
ro C
QJ
^ ro
to o» 3
Q) >>
o c
■O r— 3
M-
>>
XI
3
p—
O
O
i- +->
■H
ro -p-
C CD O
o
tsJ
+->
QJ
1— SZ
•1—
O
•r- l-
DlT-
QJ
E O
to +-»
C71
QJ X3
z
>» s-
a>
>)
cn +->
C T3
XI
O +->
3 ro
i- 'P-
f— ■»->
^ • E
+->
♦f—
•p- 'P-
to
-o a.
»/1
3 O
^ to
(_> <U 3
•r—
x:
to
Nl X
x:
QJ
'1—
S-
+-> ro
cn
fO -M C
to
•r~
•r- O
■e
QJ 3
> +J
QJ
p—
o
•r- +J
<D ro
J_
■o
3
S- TD
•1- c
x:
3 QJ
^ c
t- to
a>
>)
>>
•r- «4-
o
3
T3 ro
■M
O E
00
x: OJ E
>
+J
X O
4->
C P-
O
QJ
1
C l-
■M 14- O
•p-
O
cn
1— Q.
H-l Q
QJ i-
m
•1- QJ
•r- -r- +J
X3
to
1 Q>
3
QJ
i- +->
3 ^ ro
ro
S_
(/) r-
E
o 3
O
3 X
vO
o
O T-
0)
i-
<u
X>
3
o
• -M
•p-
O. QJ
o +->
>» J- Q
■o
■•->
c
ro
E
c p-
to
x:
□c
• 1
-H Q.
•p-
3
0)
• O.
•f—
•r—
QJ 4-*
3
SZ QJ
Q. LD "O
•1—
3
0)
cn
ro ro
+->
x:
+-> C
•<- +->
•r-
•o -»-» •
c
•1- O
o.
-H
ro ro
to
ro
+->
•r- o un
+->
i-
I.
o
QJ •!-
+->
■H
to i_
to
Q. to
O C
ro
l-
(U
QJ 0)
cn 3
■H i-
ro
QJ QJ
ro :r
ro
+-» i-
SZ
c
to O
+->
•f— 1 —
•!->
3 QJ
r— Q. +-»
(U
o ro
+->
ro to
X2 Q
•r—
u o
cz
o E
>1i—
c
to
c
ro
ro
•p-
S- +->
3 E XI
QJ +->
fO
r- O
XI .r- 3
Qi
E -Q to
3
ro
U1 •!-
to
Q.
H
"O
Qi «/^
3 o
to
ro
c
n: +->
SZ
to +->
1/)
c
TD 1—
o;
+->
Q.
>» to
to
Ol'C-
to QJ
O
E
<u E a>
•r-
OJ
c -o
c
-Q
3 XI
o
QJ C
«/)
> •
C to XI
O.X)
E ^
ro
o
Q) ro
O
•p-
E
'i~
to
-H +->
•r- T-
rO
3
+J
E X
•p-
-o x:
QJ >»-a
O >»
c
•r- C
E c +J
OJ
x:
E
o
+-»
<_) fO
•1—
to ro
(0
-H
> 3
S- ro C
C T-
•p-
3 QJ
ro E
o
E
tJt
to
•r- O
fO
TD
-H
c
•p- ^
T3
c
ra
• •
O
■M O
+-> S- to
•p-
O.
o
ro -M
C
J- ro
O 1
:xz QJ
o
E
O
(U o xt
•a
O
O
•r-
+->
o
QJ C
X3
SZ
Q. E
o
lO 0)
■O ro
c
ro
■»->
C to
o
> -p-
S- to
•r-
o
i-
• •> 1—
sz
ro
1
ro
O 3
O E
ro C
O to
o
O
-O X3
c o
c=
i-
O r—
u
o
u o
■»-»
't—
c to
O •!- to
3=
•p-
ro
-M
•p-
3
•p-
to
■o
E
LD
•f- x: c
O.
•f—
C
XJ
o
I- -H
QJ
>> c
V
>
+-» 3 o
ro
i_
0)
r— O
•f—
i- '1-
O ro
XI
-M ro
r—
zn.
•p- 'P-
o
(U
o
•r- ro
o
•1 —
•p-
Q
f «
to -o +->
O
-f->
c
+-> XJ
ro
c
■H
to
> u
O
O C T-
<_>
o
to O
3
to c
•P- T-
+-»
O -M
o. o -o
0)
M-
ro
<_)
•1—
c
QJ U-
<U Q>
■H M-
i- O
E c
>
X>
o x:
•r-
+J O
O
•1— '1—
c
cn +J
o <u o
•p-
p—
+
• h-
QJ C
<4-
to (_)
O
I/O
U ^ (->
-H
lO
<->
ro
QJ
-H •
O O
i_
cz QJ
<U
I- «
•1—
O)
•p-
>
>
ro O
>> U
Q>
QJ O.
O 1—
O) O) <u
to
r—
x:
Q)
r-
•r—
»/i
O.
to to
Q.
+->
O V
> CD c:
c=
to
CLx:
Q. i-
i_
I- CO
1 +
■»->
•r-
Q.
•r- C -P-
a>
o
+->
rO
-C
QJ
QJ
-O SZ
o
1/1
a>
+-> to 1—
to
0)
i_
+-> r—
1—
> 1
"p-
ro
Q) .f-
O
(U
C
+-> to
ro i- ro
+->
+->
L.
ro 3
+-» x:
O
^ ro
u-
(U
•r— T-
to
ro
o
<D
O
>>o
U CD.f-
4_ i_
-H
u-
-o
•r- ro 1—
ro
tz
i_
-M
to -p-
p— •
ro
to
ro
o
UJ
•1- CL
r— ro
•r-
<u
ro
r— +->
o
x: c\j
x:
to
E to
Oi
3 x: E
ro -U
•H
E
+->
Qi
•r- l-
C7J
■o
ro
M-
o
X o
3 C l-
O
o
0)
l_
O ro
<N
•1- :e
c o
i-
O to
y-
_j
<U (_>
O ro O
c
T3
3C
CO
31 o.
ro +->
X)
Z 'P-
to
QJ
+->
QJ
U
>>
E
o
c:
147
However, as was discussed in a previous section, acidic deposition tends to cause cation
leaching from vegetation which may in fact ameliorate to some extent any acidifying
processes by increasing the soil buffering capacity.
Soil litter decomposition appears to be retarded only if plant residues are
treated with extremely acidic simulated precipitation (pH 2.0) or fumigated at concen-
tration of sulphur dioxide as high as 530 ppb (Visser et al . 1987). Litter decay does
not appear to be affected by acidic precipitation of pH 3.0 to 3.5 based on the results
of both field and laboratory studies reviewed by Visser et al . (1987). In studies which
did demonstrate reductions in the rate of decay, such effects were observed at a variety
of pH values; for example, pH 1.8, and 3.7 to 4.1 for pine, pH 2.9 for spruce needles;
and pH 3.7 to 5.0 for birch leaves. In field studies where the input of acidic deposi-
tion could not be controlled, reductions of pH from 4.8 to 3.3 or to 3.9 in forest
litter layers resulted in an overall increase in litter mass due to retardation of the
decomposition process. This suggests that chronic exposures may affect microbial
decomposition processes more than acute exposures used in most laboratory studies
reviewed by Visser et al. (1987).
The studies performed to date suggest that the activity of most soil enzymes
are adversely affected by simulated acidic precipitation but only under conditions of
pH 2.0 or less.
Visser et al. (1987) also cited studies which indicated that certain portions
of the nitrogen cycle could be adversely affected by acidic deposition. This was found
to be particularly true of nitrification and nitrogen fixation which were shown to be
inhibited in soils below a pH of 6.0. Bacteria such as Rhizobium would be particularly
vulnerable under such conditions.
The effect of acidic deposition on plant-pathogen interactions appears to be
small under chronic conditions. Although studies on the effects of acidic emissions on
mycorrhizal activity are few, they indicate that these symbiontic relationships are also
largely resistant to chronic levels of acidic deposition. Visser et al. (1987) pointed
out that most ectomycorrhi zal plants are also associated with a diverse array of acid
resistant fungi which likely provide a strong defense against environmental change
caused by acidic deposition. Certain VA mycorrhizal associations, however, are acid
sensitive. Reductions of as little as 0.5 to 1.0 pH unit may render these mycorrhizal
associations non-functional, resulting in reduced plant growth.
For additional details the reader is referred to Visser et al. (1987). Many of
the studies reviewed were primarily conducted in the laboratory under extremely unreal
conditions of acidity and in pure cultures. Only in a few instances were realistic
field experiments conducted. As Visser et al. (1987) pointed out, more research is
required under field conditions and ambient concentrations to clarify the mechanisms and
dosage levels that should be of concern.
6.3 ACIDIC DEPOSITION AND INORGANIC SULPHUR MICROBIOLOGY
The previous section of this report dealt with microbiology as a whole. This
section deals solely with inorganic sulphur microbiology based on the review of the
subject by Laishley and Bryant (1987).
148
Sulphur deficiencies in various agricultural soils throughout the world have
stimulated interest in soil sulphur systems (Coleman 1966; Wainwright 1978; and Beaton
and Soper 1986). Sulphur as sulphate is required by plants for the synthesis of sulphur
containing amino acids and eventually protein synthesis. In humid regions of the world,
most of the soil sulphur is in the biologically unavailable organic form. Studies by
Walker and Adams (1958), Freney (1961), and Tabatabai and Bremner (1972) have shown that
organic sulphur represents between 42% and 93% of the total sulphur in many soils in
the world. The breakdown of organic sulphur into its biologically useable sulphate form
is accomplished by two stages: mineralization of organics; and by transformation of the
resulting inorganic sulphur to sulphate. These processes are accomplished by hetero-
trophic microorganisms which derive their energy from the decomposition of organic
matter. If the sulphur content of the substrate is greater than what can be utilized in
biosynthesis by the microorganisms, the excess sulphur is made available for other soil
processes (Reuss 1975). The global sulphur cycle is shown in Figure 4, showing the key
microbial groups responsible for sulphur cycling.
6.3.1 Oxidation Reactions
Inorganic sulphur oxidation and reduction can occur under aerobic or anaerobic
conditions and with or without light (Jorgensen 1982). Three major groups of micro-
organisms are important in the rapid oxidation of soil sulphur. These are as described
by Keunen (1975) and Wainwright (1978):
1. The colourless sulphur bacteria of the following families: Thio-
bacteriaceae, Beggiatoaceae, and Achromatiaceae;
2. Photosynthetic S-bacteria of the Chromatiaceae and Chlorobacteriaceae ; and
3. Heterotrophic microorganisms which do not gain energy from the oxidation
of sulphur compounds, actinomycetes , bacteria, and fungi.
Wainwright (1978) stated that among these groups, the colourless sulphur
bacteria and the heterotrophic microorganisms are the most important in cycling soil
sulphur. The photosynthetic bacteria of Group 2 are important in aquatic systems and/or
in flooded soils.
Laishley and Bryant (1987) di scussed* the physiological requirements of colour-
less sulphur bacteria principally of the genus Thiobaci 1 lus , its method of obtaining
energy from the breakdown of sulphur, and its acid tolerance or sensitivity. The chemi-
cal reactions by which various species of the Thiobaci 1 lus break down sulphur products
are shown in Table 33. Laishley and Bryant (1987) also described the environmental
requirements of three genera of sulphur bacteria found in hot springs (Sulfolobus,
Thiomicrospora, and Themothrix) . All of these bacterial groups thrive in areas rich in
hydrogen sulphide. The genus Beggiatoa also thrives in the presence of high hydrogen
sulphide concentrations and has been found in sulphur springs, mid layers of lakes, and
in water polluted with sewage (Laishley and Bryant 1987).
The fourth group of bacteria described by Laishley and Bryant (1987) are the
phototrophic sulphur bacteria of the Chromatiaceae and Chlorobiaceae. In contrast to
plants, these bacteria photosynthesize under anoxic conditions, use hydrogen sulphide as
149
150
Table 33. Chemical reactions of the thiobacini.
(1) 2S° + 302 + 2H2O
(2) NaaSaOa + 2O2 + H2O
(3) 2Na2S406 + 7O2 + 6H2O
(4) 2KSCN + 4O2 + 4H2O
(5) 5S + 6KN03 + 2H2O
(6) 5Na2S203 + 8NaN03 + H2O
(7) 12FeS04 + 3O2 + 6H2O
Bacterium
denitrif icans
T. thioparus
J_,_ thiooxidans
T. f errooxidans
T. novel lus
2H2SO4
^ Na2S04 + H2SO4
-> 2Na2S04 + 6H2SO4
-> (NH4)2S04 + K2SO4 + 2CO2
-> 3K2SO4 + 2H2SO4 + 3N2
9Na2S04 + H2SO4 + 4N2
-> 4Fe2(S04) + 4Fe(0H)3
Reaction(s) carried out
1. 2, 3, 4, 5, 6
1. 2, 3. 4
1, 2, 3.
1. 2. 7
2. 3
Source: Starkey (1966).
151
an electron donor, and produce elemental S, inside or outside the cell. In anaerobic
aquatic environments these bacteria oxidize hydrogen sulphide and elemental sulphur to
sulphate and in the process derive energy (ATP) required for carbon dioxide fixation.
These phototrophic bacteria are mainly restricted to the aquatic systems within a fairly
narrow environmental range defined by their requirement for light, hydrogen sulphide,
and low oxygen concentrations. These organisms are also further restricted by their
tolerance to hydrogen sulphide. Although their importance in sulphur cycling is
primarily related to aquatic systems, they are also thought to be of potential importance
in flooded soils (Laishley and Bryant 1987).
6.3.2 Heterotrophic Microorganisms
Heterotrophic microorganisms which include bacteria, fungi, and actinomycetes
are capable of oxidizing reduced forms of inorganic sulphur. Bacteria such as Arthro-
bacter, Bacillus, and Flavobacterium oxidize elemental sulphur and bisulphite to
sulphate; other species such as Achromobacter sp. and Pseudomonas sp. have been found to
oxidize S° and SsOa^" to SaOs^ (Wainwright 1979). Few reports are available that docu-
ment the sulphur oxidation properties of fungi and actinomycetes (Wainwright 1979;
Germida et al. 1985). Wainwright (1978) found, however, that the fungus Penicillium
decumbens was capable of oxidizing elemental sulphur to bisulphite.
6.3.3 Reduction Reactions
Under anaerobic conditions, and at neutral pH, the sulphate reducing bacteria
have the unique ability to utilize S04^ as an electron acceptor. The process,
referred to as dissimi latory sulphate reduction, produces copious amounts of hydrogen
sulphide in nature and may also be involved in many geochemical phenomena (Peck 1961,
1975; LeGall and Postgate 1973). It should be noted that another process, assimilatory
sulphate reduction, is also common to most plants and bacteria and involves the reduction
of just enough S04^ to HS to meet cellular requirements for sulphur containing
amino acid biosynthesis (Peck 1961, 1975).
There are two main groups of sulphate reducing bacteria. Group one consists of
Desulfovibrio, Desulfomonas , and Desul f otomucul urn. These genera utilize lactate and
occasionally pyruvate or ethanol as carbon sources. The second group consists of
Desulfobulbus , Desul fobacter, Desul fococcus . Desulfosarcina, and Desulfonema. These
genera, in contrast to those of group one, utilize the oxidation of the fatty acid
acetate to derive their energy.
An example of the reaction pathways of Desulfovibrio in the oxidation of lactic
acid and the dissimi latory reduction of sulphate is shown in Figure 5 (Laishley and
Bryant 1987).
Also in the context of nonclassical sulphate reduction, some species of the
genera Salmonella . Proteus, Camphylobacter. Succhuromyces . and Pseudomonas have been
shown, in pure culture, to anaerobical ly reduce small amounts of SOa^" to HS~
(McCready et al. 1974; Brock et al. 1984). The role these organisms play in the cycling
of inorganic sulphur in the ecosystem is not known at this time (Laishley and Bryant
1987) .
152
H
I
2CHpC- COOH
I
OH
LACTIC ACID
GROWTH
V
■►2XH2
2 CH3C - COOH
r
CO.
2CoA
//
CH,-C - Co A
2Pi
2CH3C - P
2CoA
AMP
2ADP-^
Substrate level phosphorylation
2 ATP
2S0
PPi
2Pi
2CH3-COOH
S04=
ACETATE
DISSIMILATORY
SO4 REDUCTION
Fd = Ferrodoxin
ETS = Electron Transport System
Figure 5. Oxidation of lactic acid and the di ssimi latory reduction of
sulphate by Desulfovibrio sp.
153
6.4 ECOLOGICAL AND ECONOMIC EFFECTS OF MICROBIAL INORGANIC SULPHUR OXIDATION AND
REDUCTION
Laishley and Bryant (1987), in their review of sulphur microbiology, have
discussed the ecological role of these microorganisms and their economic importance.
Only the most important aspects will be discussed here.
As stated previously, Thiobaci 1 1 us species oxidize elemental sulphur and in the
process, produce sulphuric acid which can cause soil acidification. As long as the
amount of sulphur available for oxidization is limited, this increase in soil acidity
can benefit plant growth by providing not only more sulphate, but also more micro-
nutrients because in the process of acidification, Ca, Mg, P, and K are also released
(Gorham 1976; Wainwright 1978). Other advantages of sulphur oxidation are an increase
in fertility in basic soils and protection from diseases such as potato blight (Starkey
1966). The addition of controlled amounts of elemental sulphur as a method of improving
soil fertility in soils with pH 8 to 9 has been suggested and tried experimentally
(Adamczyk- Winiarska et al. 1975; Bollen 1977; and Legge et al . 1986). The results of
the tests showed a rapid reduction in soil pH as a result of the additions.
If large amounts of acid are added to a soil system the resultant lowering of
the pH can adversely affect the soil microflora. A reduction of heterotrophic bacterial
populations is one effect of such inputs ( Adamczyk-Wi niarska et al. 1975; Wood 1975;
Bryant et al . 1979; and Wainwright 1979). Bryant et al. (1979) have also shown that the
respiratory activity of microflora responsible for the degradation of glucose, starch,
cellulose, casein, and urea was significantly reduced in soils exposed to simulated
acidic rain of pH 3.0. Tamm (1976) reported that acidification inhibited the nitrifi-
cation process in a forest soil while other workers have demonstrated that nitrogen
fixation was inhibited under acidic conditions (Oden 1971; Dochinger and Seliga 1975).
In addition, Tamm (1976) suggested that mycorrhizal fungi are very sensitive to acidic
conditions. However, as noted earlier, most fungi are not acid sensitive.
6.4.1 Oxidation of Metal Sulphides in Soil
A number of environmental and agricultural problems occur as a result of
chemolithotrophic oxidation of pyrite (FeS2), a compound commonly present in coal and
coal mining effluents. Pyrite is also found in soils where hydrogen sulphide from
bacterial reduction reacts with ferrous iron to form ferrous sulphide. This in turn can
react with either elemental sulphur or sulphides to form pyrite (Metson et al. 1977).
Pyritic soils are potential sites for acidification. If sites containing pyritic soils
change from anaerobic to aerobic conditions, acidic soils called "cat clay" will be
formed (Metson et al. 1977). Under aerobic conditions pyrite is oxidized to sulphuric
acid, jarosite [KFe3(S04 ) 2(0H) e] , ferric oxides, and ferric sulphate (Bloomfield
and Coulter 1973; Metson et al . 1977; and Kargi 1982). Pyrite may also be subject to
oxidation by certain species of Thiobaci 1 lus . There are two basic mechanisms by which
pyrite is oxidized; these are the direct and the indirect methods as shown in Figure 6.
All of the reactions in Figure 6 utilize Thiobaci 1 1 us f erroxidans to complete the
oxidation .
154
DIRECT AND INDIRECT
OXIDATION MECHANISMS FOR PYRITE OXIDATION
DIRECT
T. ferrooxidans
(1) 2 Fe $2+ HjO + 7.5 ► Feg ($04)+ H2SO4
0
Pyrite Ferric sulfate
INDIRECT
T. ferrooxidans
(2) 2FeS2+ H2O +7.5O2 ► Feg (SO4) + H2SO4
0
. . chemical
(3) FeSp+ FeplSOj, ►3FeS04+ 2S0
^ ^ ^3 oxidation ^
o T. ferrooxidans
(4) 2S^ + 3O2+ 2H2O ►2H2SO4
T. ferrooxidans
(5) 4 Fe SO4 + 2 H2SO4+ O2— — ► 2 ?^z^SO^\ + ^H^O
Figure 6. Direct and indirect oxidation mechanisms for pyrite oxidation.
155
Other acidophilic thiobacilli may also participate in the oxidation of the
elemental sulphur produced in reaction [3] in Figure 6 and thereby catalize the rapid
formation of sulphuric acid. One of the more serious economic problems associated with
sulphur microbiology results from this process and consists of the corrosion of metal,
and weathering of stone and concrete (Parker 1947; Bryant et al . 1985).
Bacterial oxidation can also be beneficial. Pyrite oxidation has been used to
extract sulphur from coal prior to combustion, thus lowering the emissions of sulphur
oxides and the potential for acidic rain (Kargi 1982).
Thiobaci 1 lus f erroxidans is also used to leach copper from low grade ores. The
process is similar to metal mobilization in soils. The bacteria grow by oxidizing
ferrous iron to ferric iron. In the process they create a strongly oxidizing acid
solution which in turn solubilizes and mobilizes other metals. Copper can then be
reclaimed where other mining techniques would likely prove uneconomical (Brierley 1978).
6.4.2 Phototrophic Sulphur Bacteria
The photosynthetic sulphur bacteria provide the sulphur reducing bacteria with
a substrate for growth without the consumption of molecular oxygen (Pfennig 1975;
Postgate 1982). They also provide a food source for protozoa in lakes, thereby
contributing to the secondary productivity of lakes (Sorokin 1970; Pfennig 1975).
In polluted waters the photosynthetic sulphur bacteria produce an excess of
hydrogen sulphide causing in turn phototrophic blooms which, because of their colourful
nature, have been suggested as indicators of water pollution (Postgate 1982).
6.5 FACTORS AFFECTING THE MICROBIAL OXIDATION OF SULPHUR
There are three main factors which influence the oxidation of elemental sulphur:
the sulphur itself; the sulphur oxidizing microorganisms; and the nature of the soil
environment where sulphur oxidation is occurring (Weir 1975).
6.5.1 Sulphur
Sulphur is a very complex and non-homogenous element. In studies conducted on
"manufactured" sulphur, Laishley et al. (1984) produced three classes of sulphur by:
(1) purifying production grade sulphur (B and F ) ; (2) rapid cooling or quenching of
molten B and F sulphur to form MMS (Mixed Molecular Sulphur); and (3) extraction of
polymeric sulphur from sulphur using CS2.
Laishley et al. (1984) showed that the B and F sulphur and the polymeric sulphur
were oxidized by Thiobaci 1 lus albertis at similar rates while the MMS was oxidized at a
much slower rate. It was clearly shown that the rate curves for these sulphur species
began to diverge only after some 5% of the total sulphur was consumed (3 days). However,
the MMS contained orthorhombic crystalline sulphur and polymeric sulphur well in excess
of this percentage, indicating that the effect of the different molecular species in MMS
was not simply related to the amount present. It has been suggested that the tightness
with which the sulphur lattice is packaged could reduce the numbers of sterically
favourable binding sites of T. albertis resulting in lower oxidation rates under certain
conditions .
156
Laishley et al. (1983) have also shown that particle size determines the total
amount of S that can be converted to sulphuric acid. Specifically, they found in their
studies that T. al berti s was capable of metabolizing 70% of a powdered sulphur with a
particle size range of 150 ym to 250 ym, while only 3% of an equivalent weight of
sulphur prill with a particle size range of 1.68 to 2.00 mm was metabolized. This type
of information is critical from both a storage and shipping point of view for industries
such as the sour gas processors. This indicates that large sized particles of sulphur
will produce less acid and will also be more efficient for shipping because more product
will reach its destination without being metabolized. Laishley et al . (1983) established
the relationship that the microbial oxidation rate of sulphur was a function of the
surface area per weight of sulphur.
Sulphur oxidizing microorganisms such as Thiobaci 1 lus have been shown to
develop what are suspected to be acidic mucopolysaccharide polymer containing threadlike
structures termed a "glycocalyx" which are utilized in substrate attachment (Takakawa
et al. 1979; Costerton and Irvin 1981; Ladd 1982; and Bryant et al. 1983,1984). Bryant
et al. (1983) believe that bacterial cells attached to the sulphur substrate by their
glycocalyx produce microcolonies and eventually biofilms within the first few days of
oxidation. This hypothesis was supported by Laishley et al. (1983) who observed with
electromicroscopy cell surface processes and growth on prill sulphur, In comparison to
powdered sulphur (70% oxidized) only 3% of the prill sulphur was oxidized over compar-
able time period. Laishley et al. (1983) also found that sulphate production remained
linear over the duration of their experiment indicating that no exponential bacterial
growth occurred as one would expect if the total substrate was available for metabolism
by the oxidizing bacteria.
Based on the aforementioned experiments, Laishley and Bryant (1987) have
proposed recommendations reducing environmental impacts caused by elemental sulphur via
microbial oxidation. They suggested that elemental sulphur be stored in the solid block
form and that production of powdered sulphur be minimized. Solid sulphur blocks would
present a limited surface area to weight ratio which in turn would limit oxidation
potential. This would increase the likelihood of biofilm development and would effec-
tively insulate the block from further degradation.
6.5.2 Soil Environment and Its Effects on Sulphur Microbiology
Microbial oxidation of sulphur is influenced by temperature. Oxidation can
occur below 10°C, although the rate will be slow (Weir 1975). Maximum rates of microbial
sulphur oxidation have been reported at 40°C (Li and Caldwell 1966). Bryant et al. (1985)
have shown that Thiobaci 1 lus albertis , a newly characterized acidophilic sulphur oxi-
dizer, is non-functional at 5° and 37°C, and its optimum functional temperature is 28°C.
At 37°C the bacterial activity is stopped and the cells are killed, whereas at 5°C the
cells are not killed. These findings are significant when one considers the amounts of
elemental sulphur stored in temperate climates where winter temperatures often reach the
lower range but seldom exceed the upper range of tolerance. The experiments conducted
by Bryant et al. (1985) showed that bacteria that were non-functional at 5°C could
be revived by raising the temperature.
157
Soil type also influences the rate at which elemental sulphur can be oxidized.
Laishley and Bryant (1987) cite the results of soil tests in Alberta that used elemental
sulphur, sulphur concrete, and Portland cement in various locations throughout the
province. After several years, thiobacilli were detected around the different sulphur
substrates even though they were basically undetectable in some cases at the start of
the experiment. A typical succession pattern was observed whereby the less acidophilic
thiobacilli created an acid environment for the more acidophilic types that followed.
Depending on the buffering capacity of the soil surrounding the test cylinders, the time
required to detect pH changes ranged from 5 years at the most sensitive site, a forest
soil, to no change on highly buffered alpine soil even after eight years of testing.
Sulphur oxidizing microorganisms are strongly affected by moisture conditions.
Often in soils, moisture status is closely linked with the oxygen status. It is known
that the thiobacilli will not grow in water-logged soils. Moser and Olson (1953)
showed, however, that moisture levels near field capacity generated maximum microbial
oxidation. Similarly, Laishley and Bryant (1985) observed that at moisture contents of
<2% in a sandy loam site, the sulphur oxidizing activity of thiobacilli was low.
Although the site contained sulphur substrates, the occurrence of these organisms was
only sporadic over the eight years of study.
158
6.6 EFFECTS OF ACIDIC DEPOSITION ON SOIL MICROORGANISMS AND MICROBIALLY MEDIATED
PROCESSES: LITERATURE CITED
Adamczyk-Winiarska, Z., M. Krol, and J. Kobus. 1975. Microbial oxidation of elemental
sulphur in brown soil. Plant and Soil 43: 95-100.
Alexander, M. 1980. Effects of acidity on microorganisms and microbial processes in soil.
In: Effects of Acid Precipitation on Terrestrial Ecosystems, eds.
T.C. Hutcinson and M. Havas. New York: Plenum Press, pp. 363-374.
Alexander, M. 1977. Introdution to Soil Microbiology. New York: John Wiley and Sons.
467 pp.
Anderson, J.P.E. and K.H. Domsch. 1978. A physiological method for the quantitative
measurement of microbial biomass in soils. Soil Biology and Biochemistry 10:
215-221 .
Babich, H. and G. Stotzky. 1978. Atmospheric sulphur compounds and microbes. Environ-
mental Research 15: 513-531.
Beaton, J.D. and R.J. Soper. 1986. Plant response to sulphur in western Canada. American
Society of Agronomy Monograph (in press).
Bloomfield, C. and J.K. Coulter. 1973. Genesis and mangement of acid sulphate soils.
Advances in Agronomy 25: 265-326.
Bollen, W.B. 1977. Sulfur oxidation and respiration in 54 year old soil samples. Soil
Biology and Biochemistry 9: 405-410.
Brierley, C.L. 1978. Bacterial leaching. CRC Critical Reviews in Microbiology 6: 207-262.
Brock, T.D., D.W. Smith, and M.T. Madigan. 1984. Biology of Microorganisms. 4th Ed.
Englewood Cliffs, New Jersey: Prentice Hall. pp. 704-715.
Bryant, R.D., E.J. Laishley, C.L. Labine, and J.B. Hyne. 1985. Thiobacilli and wet sul-
phur corrosion. ASR Quarterly Bulletin XXII: 1-11.
Bryant, R.D., J.W. Costerton, and E.J. Laishley. 1984. The role of Thiobacillus albertis
in the adhesion of cells to elemental sulfur. Canadian Journal of Microbiology
30: 81-90.
Bryant, R.D., K.M. McGroarty, J.W. Costerton, and E.J. Laishley. 1983. Isolation and
characterization of a new acidophilic Thiobaci llus species (T. albertis) .
Canadian Journal of Microbiology 29: 1159-1170.
Bryant, R.D., E.A. Gordy, and E.J. Laishley. 1979. Effect of soil acidification on the
soil microflora. Water, Air, and Soil Pollution 11: 437-445.
Coleman, R. 1966. The importance of sulphur as a plant nutrient in world crop production.
Soil Science 101 : 230-239.
Costerton, J.W. and R.T. Irvin. 1981. The bacterial glycocalyx in nature and disease.
Annual Review of Microbiology 351: 299-324.
Dochinger, L.S. and T.A. Seliga. 1975. Acid precipitation and the forest ecosystem.
Journal of the Air Pollution Control Association 25: 1103-1105.
Freney, J.R. 1961. Some observations on the nature of organic compounds in soil. Aus-
tralian Journal of Agricultural Research 12: 424-432.
Germida, J.J., J.R. Lawrence, and V.V.S.R. Gupta. 1985. Microbial oxidation of sulphur
in Saskatchewan soils. In: Proceedings of Sulphur - 84, an International
Conference on Sulphur, Calgary, Alberta. 1984 June 3-6; Sulphur Development
Institute of Canada; pp. 703-710.
Gorham, E. 1976. Acid precipitation and its influence upon aquatic ecosystems - an over-
view. Water, Air, and Soil Pollution 6: 471-481.
159
Jorgensen, B.B. 1982. Ecology of the bacteria of the sulphur cycle with special reference
to anoxic-oxic interface environments. Philosophical Transactions of the Royal
Society of London B 298: 543-561.
Kargi, F. 1982. Microbiological coal desulphuri zation . Enzyme Microbiology and Tech-
nology 4: 13-19.
Kellogg, W.W., R.D. Cadle, E.R. Allen, A.L. Lazrus, and E.A. Martell. 1972. The Sulfur
Cycle. Science 175: 587-596.
Keunen, J.G. 1975. Colorless sulfur bacteria and their role in the sulfur cycle. Plant
and Soil 43: 49-76.
Ladd, T.I. 1982. Heterotrophic activity of epilithic bacteria communities in a lotic
ecosystem. Ph.D. Thesis. Calgary, Alberta: University of Calgary. 402 pp.
Laishly, E.J., M.G. Tyler, and R.G. McCready. 1978. Environmental assessment of soils in
contact with sulfur-based construction material. Iri: Environmental
Biogeochemistry and Geomicrobiology . Volume 2: The Terrestrial Environment,
ed., E. Kbenbein, Ann Arbour, Michigan: Ann Arbor Science, pp. 699-705.
Laishley, E.J. and R. Bryant. 1987. Critical Review of Inorganic Sulphur Microbiology
with Particular Reference to Alberta Soils. Prep for the Acid Deposition
Research Program by the Department of Biology, The University of Calgary.
ADRP-B -04-87. 56 pp.
Laishley, E.J. and R.D. Bryant. 1985. Long term environmental effects of sulphur and
sulphur concrete on the soil ecosystem. In: Proceedings of Sulphur-84, an
International Conference on Sulphur, Calgary, Alberta. 1984 June 3-6; Sulphur
Institute of Canada; pp. 641-646.
Laishley, E.J., M.G. Tyler, and H.R. Krouse. 1984. Sulfur isotope fractionation during
SOa^- reduction by different clostridial species. Canadian Journal of
Microbiology 30: 841-844.
Laishley, E.J., R.D. Bryant, B.W. Kobryn, and J.B. Hyne. 1983. The effect of particle
size and molecular composition of elemental sulphur on ease of microbiological
oxidation. Alberta Sulphur Research Ltd. Quarterly Bulletin XX(3): 33-50.
LeGall, J. and J.R. Postgate. 1973. The physiology of the sulfate reducing bacteria.
Advances in Microbial Physiology 10: 81-133.
Legge, A.H., J. Corbin, J. Bogner, M. Strosher, D. Parkinson, H.R. Krouse, E.J. Laishley,
M.J. Cavey, C.E. Prescott, R.J.F. Bewley, M. Nosal, H.U. Schellhase, T.C.
Weidensaul, and J. Mayo. 1986. Acidification in a temperate forest ecosystem:
The role of sulphur gas emissions and sulphur dust. A Final Report Prepared
for the Whitecourt Enviromental Study Group by Kananaskis Centre for Environ-
mental Research, The University of Calgary, Calgary, Alberta. 471 pp.
Li, P. and A.C. Caldwell. 1966. The oxidation of elemental sulfur in soil.
Proceedings of the Soil Science Society of America 30: 370-372.
McCready, R.G.L., E.J. Laishley, and H.R. Krouse. 1974. The effect of sulfate and sul-
fite on growth and sulphur isotope fractionation by Clostridi urn pasteurianum.
Canadian Sulphur Symposium. 1974 May 30-June 1; The University of Calgary,
Calgary, Alberta; pp. F1-F7.
Metson, A.J., D.J. Gibson, J.E. Cos, and D.B. Gibbs. 1977. The problem of acid sulphate
soils, with examples from North Auckland, New Zealand. New Zealand Journal of
Science 20: 371-395.
Moser, U.S. and R.V. Olson. 1953. Sulfur oxidation in four soils as influenced by soil
moisture tension and sulfur bacteria. Soil Science 76(4): 251-257.
Oden, S. 1971. Some effects of the acidity of air and precipitation. Proceedings of
International Air Pollution Control and Noise Abatement Exhibition Conference
1: 85-94.
160
Parker, CD. 1947. Species of sulphur bacteria associated with corrosion of concrete.
Nature 159: 439-440.
Peck, H.D., Jr. 1975. The microbial sulfur cycle in sulfur in the environment. In: Sulfur
in the Environment, ed . H.S. Parker. St. Louis, Missouri: Missouri Botanical
Gardens and Union Electric Company, pp. 62--79.
Peck, H.D., Jr. 1961. Enzymatic basis for ass imi ' atory and dissimatory sulfite reduction.
Journal of Bacteriology 82: 933-939.
Persson, T., E. Booth, M. Clarholm, H. Lundvist, B.E. Soderstrom and B. Sohlenius. 1980.
Trophic structure, biomass dynamics and carbon metabolism of soil organisms in
a Scots pine forest. Xn: Structure and function of northern confierous forests
- an ecosystem study, ed . T. Persson. Ecological Bulletin (Stockholm) 32:
419-459.
Pfennig, N. 1975. The phototrophic bacteria and their role in the sulfur cycle. Plant
and Soi 1 43: 1-6.
Postgate, J.R. 1982. Economic importance of sulphur bacteria. Philosophical Transac-
tions of the Royal Society London B 298: 583-600.
Reuss, J.O. 1 975. Sulfur in the soil system. Ijn: Sulfur in the Environment, ed . H. S.
Parker. St. Louis, Missouri: Missouri Botanical Gardens and Union Electric
Company, pp. 51-62.
Sorokin, Yu. I. 1970. Interrelations between sulphur and carbon turnover in meromictic
lakes. Archiv Hydrobiologia 66: 391-466.
Starkey, R.L. 1966. Oxidation and reduction of sulfur compounds in soils. Soil Science
101(4): 297-306.
Tabatabai, M.A. and J.M. Bremner. 1972. Forms of sulfur and carbon, nitrogen and sulfur
relationships in Iowa soils. Soil Science 114: 380-386.
Takakawa, S., T. Fujimori, and H. Iwasaki. 1979. Some properties of cell-sulfur adhesion
in Thiobaci 1 lus thiooxidans . Journal of General and Applied Microbiology 25:
21-29.
Tamm, CO. 1 976. Acid precipitation. Biological effects in soil and on forest vegeta-
tion. Ambio 5: 235-238.
Wainwright, M. 1979. Microbial S-oxidation in soils exposed to heavy atmospheric pollu-
tion. Soil Biology and Biochemistry II: 95-98.
Wainwright, M. 1978. Microbial sulphur oxidation in soil. Science Progress Oxford 65:
459-475.
Walker, T.W. and A.F.R. Adams. 1958. Studies in soil organic matter. Soil Science 85:
307-318.
Weir, R.G. 1975. The oxidation of elemental sulphur and sulphides in soil. In: Sulphur
in Australasian Agriculture, ed . K.D. McLachlan. Sidney, Australia: Sidney
University Press, pp. 40-50.
Wood, T. 1975. Acid precipitation in sulfur in the environment. In: Sulfur in the
Environment, ed . H.S. Parker. St. Louis, Missouri: Missouri Botanical Gardens
and Union Electric Company, pp. 39-50.
161
7 . EFFECTS OF ACIDIC DEPOSITION ON GEOLOGY AND HYDROGEOLOGY
7.1 GROUNDWATER HYDROLOGY
The potential for geologic and hydrologic effects as a result of acidification
have been reviewed by Campbell (1987). Many of the processes discussed by Campbell such
as buffering capacity, cationic exchange, pH and the acidification process, metal
mobilization, and toxicity have been previously discussed in this overview and, there-
fore, only the highlights of Campbell's report will be included here. A generalized
view of the hydrologic cycle is shown in Figure 7.
The land-based portion of the hydrologic cycle is relatively well understood
and although complex, is amenable to quantification. Groundwater hydrology is a measure
of physical parameters essentially unaffected by chemical processes. Although ground-
water hydrology plays a major role in influencing groundwater chemistry, groundwater
hydrology itself is largely unaffected by geochemical processes. One can be reasonably
certain that basin hydrology will be unaffected by geochemical processes that may be
stimulated by acidic deposition. Therefore, one can conclude with confidence that the
hydrology of a basin before acidic deposition will be the same as the hydrology of the
basin after it occurs. However, the chemical constituents composing the basin waters
may be totally changed.
Scientists who concentrate on the soils portion of the subsurface have tended
to think of groundwater as a sink for soil leachate and that the problems of acidifica-
tion were nullified once this system was reached. Recently, studies have shown that
groundwater constitutes almost all of the base flow and a high percentage of peak flow
from basin runoff. This is particularly important when one considers that in much of
North America, winter flows in rivers are primarily of the base flow variety and, there-
fore, toxic metals have little opportunity to be diluted during this portion of the
year. Therefore, in order to fully understand the chemistry of surface waters, it is
essential that the groundwater chemistry also be considered. The ILWAS (Integrated Lake
Water Acidification Study) mentioned in the sections on surface waters and soils
(Sections 5 and 8 of this report), constitutes at present one of the most comprehensive
examinations of receptor responses to acidic deposition. The results of ILWAS show that
the routing of water through a watershed is the major determinant of lake water
alkalinity and vulnerability to acidification as a result of deposition processes.
These findings, therefore, dictate that any study of acidification and its effects must
consider effects on groundwater, both surficial and deep, in order to produce viable
results .
7.2 HYDROGEOLOGICAL NEUTRALIZATION PROCESSES
It has been recognized that acidic input into the land -based portion of the
hydrologic cycle can be neutralized if sufficient quantities of buffering agents are
present, and if the rate of neutralization can balance the rate of acidic input. With
time, consumption of buffering agents by the various neutralization processes may lead
to the exhaustion of the neutralizing ability of the soi 1 -rock-water system. In
attempting to quantify acidification or potential acidification, scientists have used a
wide range of chemical descriptors including pH, alkalinity, cation exchange capacity.
162
163
buffering capacity, calcite saturation index, acid neutralization capacity, and lime
potential. These descriptors are all described in detail in Sections 5 and 8 of this
report.
As many as five types of neutralization processes may be active in the subsur-
face, (Last et al. 1980), but Bache (1984) has contended that the two major processes
are cation exchange and acid hydrolysis reactions.
7.3 EVIDENCE OF GROUNDWATER ACIDIFICATION
Strong evidence of groundwater acidification has been observed in a number of
Scandinavian studies. Evidence of groundwater acidification includes declining pH
values and alkalinity as reflected in the bicarbonate values, coupled with an increase
in calcium and sulphate. In the studies of Hultberg and Johansson (1981) and Jacks and
Knutsson (1981), shallow aquifers with a low buffering capacity were most sensitive to
acidification. In the La Holm region of Holland, up to 50% of the domestic water wells
had a pH below 5.5 (Hultberg and Johansson 1981). In this study, several municipal
water supplies in deeper aquifers also showed signs of acidification.
In Canada, analyses of water samples from wells in the Sudbury, Muskoka, and
Haliburton areas of Ontario have not shown the same evidence of acidification as seen in
Sweden. In each of the Canadian studies the systems lacked buffering capacity and have
been exposed to long-term acidic deposition either from local sources and/or from long
range transport. Only a small percentage of the water wells sampled in the Ontario
studies showed depressed pH. It is conceded, however, that due to the heavy atmospheric
loading of acidic materials in these areas and because of the low buffering capacity
available for neutralization, with no change in the present loading, acidification of
the groundwater will occur given sufficient time.
7.4 EFFECTS OF ACIDIC DEPOSITION ON MAJOR CATIONS AND ANIONS IN GROUNDWATER
The chemical constituents that are involved in the acid neutralization process
are dependent upon the chemistry of the rock influenced water. The major cations and
anions are: calcium, magnesium, sodium, potassium, carbonate/bicarbonate, sulphate,
chloride, and nitrate. The role of each of these ionic species in the neutralization
process is discussed in Sections 5 and 8 of this report.
Evidence is available to suggest that the following chemical changes may be
observed upon acidification:
1. Increase in calcium concentrations;
2. Mobilization and increase in groundwater sulphate, if the adsorption
capacity of the rock material is exceeded;
3. Mobilization and movement of chloride into the water table, resulting in
increased groundwater concentrations;
4. Increase in nitrate with a net beneficial effect in nitrogen deficient
soils. Once the nitrogen saturation of the subsurface waters is exceeded,
nitrate may move to the water table. Once at the water table, no major
attenuation mechanisms have been recognized;
164
5. Decrease in pH with an accompanying reduction in bicarbonate alkality; and
6. An increase in the concentration of dissolved aluminum, possibly accom-
panied by increased mobility of other dissolved metals.
7.5 EFFECTS OF ACIDIC DEPOSITION ON METALS IN GROUNDWATER
As noted in the sections on surface water (Section 8) and soils (Section 5) in
this report, when pH levels decrease metals become increasingly mobile. This is parti-
cularly true in the subsurface between the pH values of 3.5 to 4.0. Metals of concern
in subsurface waters are: aluminum, zinc, lead, copper, manganese, arsenic, chromium,
cadmium, nickel, mercury, iron, and selenium. Because of the abundance of aluminum in
rock materials and its highly toxic nature, this metal is of particular concern and is
often used as an indication of acidification. Elevated aluminum levels have been
detected in both surface and groundwaters in Sweden at locations where atmospheric
loading has been heavy and acidification has been documented (Johansson and Hultberg
1977; Hultberg and Johansson 1978). Similar evidence has been found in studies conduc-
ted by Sharpe et al. (1984) in southern Pennsylvania. Studies conducted in eastern
Canada have not as yet detected this trend in the subsurface waters except in a few
samples. In most cases, metal concentrations were similar to background levels (Sibul
and Vallery 1982; Sibul and Reynolds 1982). It should be noted, however, that the
process of adsorption and precipitation may reduce the mobility of metals as they move
from low to high pH environments. Therefore, it could be suggested that the subsurface
waters have not as yet exhausted their buffering ability in those areas studied in
eastern Canada and that the problem has not as yet reached the proportions of that found
in Sweden.
Information currently available in the literature on the chemistry of heavy
metals in natural waters and soils show that metal solubility in aqueous systems can be
influenced by the following factors: organic materials, clay minerals, metal hydroxides,
pH, and divalent ions.
7.6 PREDICTION OF ACIDIC DEPOSITION EFFECTS ON GROUNDWATER
7.6.1 Sensitivity Analysis
Early attempts to assess sensitivity of geographic areas to acidic deposition
focussed on the ability of soil and rock materials to neutralize acidity (Shilts 1980,
cited by Saskatchewan Research Council 1982; Glass et al. 1982). A major weakness of
this approach is that it only addresses one portion of the ecological setting. It has
subsequently been recognized that sensitivity of a geographic setting can only be
assessed within the context of a multidisciplinary examination of the entire ecosystem.
The present Acidic Deposition Research Program as represented by this overview is an
attempt to take this latter approach to analysis.
7.6.2 Modelling
The development of hydrological and geochemical models is relatively well
advanced. Presently, any number of computer models exist for simulating a wide variety
of groundwater scenarios in both the saturated and unsaturated zones. In the past ten
years, groundwater hydrology models have been coupled with mass transport models to
165
allow groundwater contamination problems to be simulated. Geochemical models which
include PHREEQE, GEOCHEM, and GEOCHEM II allow the simulation of complex geochemical
scenarios. The ILWAS has developed a model that has successfully simulated a variety of
geochemical and hydrologic conditions of interest to researchers investigating the
impacts of acidic deposition.
7.6.3 Human Impacts
Evidence presented to date indicates that there are identifiable issues con-
cerning potential human health effects from acidification of groundwater. The three
major areas of concern are:
1. The acidification of groundwater leading to alteration in the distribution
of major cations and anions in domestic and municipal groundwater supplies
with ensuing impact on the potability and aesthetics of the water supply.
2. The leaching of toxic metals from watersheds and from water storage and
distribution systems.
3. The contamination of edible fish by toxic metals. Metals considered to be
of concern include: lead, mercury, aluminum, copper, zinc, cadmium,
chromium, iron, nickel, and selenium.
166
7.7 EFFECTS OF ACIDIC DEPOSITION ON GEOLOGY AND HYDROGEOLOGY : LITERATURE CITED
Bache, B.W. 1984. Soil-water interactions. Philosophical Transactions of the Royal
Society of London 8305: 135-149.
Campbell, K.W. 1987. Pollutant Exposure and Response Relationships: A Literature
Review. Prep for the Acid Deposition Research Program by Subsurface Technolo-
gies and Instrumentation Limited, Calgary, Alberta, Canada. ADRP-B-07-87 .
151 pp.
Glass, N.R., D.E. Arnold, J.N. Galloway, G.R. Hendrey, J.J. Lee, W.W. McFee, S.A. Norton,
C.F. Powers, D.L. Rambo, and C.L. Schofield. 1982. Effects of acid precipita-
tion. Environmental Science and Technology 16(3): 162A-169A.
Hultberg, H. and S. Johansson. 1981. Acid groundwater. Nordic Hydrology 12: 51-64.
Hultberg, H. and S. Johansson. 1978. Rapport Rorande Orsakerna Till NIetall-och Syra-
belastningen av Grundvattenti 1 1 ri nni ngen Till Delar av Stenunge a i Stenung-
sunds. Kommun. (Report Concerning the Causes of Metal and Acidification of
Groundwater Inflow to Parts of Stenunge Stream in the Parish of Stenungsund) .
Report from Swedish Water and Air Pollution Research Institute. Gothenburg,
pp. 12-14.
Jacks, G. and G. Och Knutsson. 1981. Kanslighet for Grundvattenf orsurni ng . Projekt
Kol-Halsa-Mi 1 jo, Statens Vattenf al 1 sverk , Rapport 11 (including an English
Summary). (Original not seen; information taken from Swedish Ministry of
Agriculture, 1982.)
Johansson, S. and H. Hultberg. 1977. Geologiska, Hydrologiska och Hydrogeologi ska
Faktorers Inverkan pa Kalkning av Forsurade Sjoar. (Geological, Hydrological
and Hydrogeological Factors Affecting Lime Treatment of Acidified Lakes).
Uppsala, Finland: Report from Division of Hydrology, University of Uppsala.
(Original not seen; information taken from Hultberg and Johansson, 1981.)
Last, F.T., G.E. Likens, B. Ullrich, and L. Walloe. 1980. Acid precipitation-progress
problems. In: Ecological Impact of Acid Precipitation, Proceedings of an
International Conference, eds. D. Drablos and A. Tollan, Sandefjord, Norway;
SNSF Project, Oslo, Norway; pp. 10-12.
Saskatchewan Research Council. 1981. Transport of acid forming emissions and potential
effects of deposition on north-eastern Alberta and northern Saskatchewan: A
problem analysis. Regina, Saskatchewan: SRC Technical Report No. 122. 43 pp.
Sharpe, W.E., D.R. DeWalle, R.T. Leibfried, R.S. DiNicola, W.G. Kimmel, and L.S. Sherwin.
1984. Causes of acidification of four streams on Laurel Hill in southwestern
Pennsylvania. Journal of Environmental Quality 13(4): 619-631.
Shilts, W.W. 1980. Sensitivity with respect to bedrock lithologies of eastern Canada.
In: Second Report of the United States-Canada Research Conciliation Group on
the Long Range Transport of Air Pollutants, eds. A. P. Altshuller and G.A.
McBean. (Original not seen; information taken from Saskatchewan Research
Council, 1981.)
Sibul, U. and D. Vallery. 1982. Acidic precipitation in Ontario Study - A synoptic survey
of the acidity of groundwaters in the Muskoka - Haliburton area of Ontario,
1982. Toronto: Hydrology and Monitoring Section, Water Resources Branch,
Ontario Ministry of the Environment APIOS Report No. 006182. 18 pp.
Sibul, U. and L. Reynolds. 1982. Acidic precipitation in Ontario study - A synoptic
survey of the acidity of groundwaters in the Sudbury area of Ontario, 1981.
Toronto: Hydrology and Monitoring Section, Water Resources Branch, Ontario
Ministry of the Environment APIOS Report No. 005182. 36 pp.
167
Swedish Ministry of Agriculture. 1982. Acidification Today and Tomorrow. Environment '82.
Committee Study. Prepared for the 1982 Stockholm Conference on the Acidifica-
tion of the Environment, trans. S. Harper. Stockholm, Sweden: 232 pp.
Universal Oil Products Co. 1972. Groundwater and Wells. Johnson Division. St. Paul,
Minnesota. 440 pp.
168
169
8. EFFECTS OF ACIDIC DEPOSITION ON SURFACE WATER ACIDIFICATION
8.1 DETERMINATION OF ACIDITY IN SURFACE WATERS
Acidity is measured by two main methods, measurement of pH and measurement of
alkalinity. Alkalinity is equivalent to the buffering capacity or increases in the
ability to neutralize H*. The major acid neutralizing species is HCOa" which
combines with to form CO2 and H2O as follows:
HCOa" + h"^ ^ CO2 (aq) + H2O [19]
Alkalinity may be defined as the concentration of all hydrogen ion acceptors minus the
free H"^ (Gherini et al. 1984). Thus:
Alkalinity = [HCOa"] + 2[C03^"] + [OH"] + [other acceptors] - [H"^]
= ANC or Acid Neutralizing Capacity [20]
In low alkalinity waters, the concentration of the other H"^ ion acceptors can become
large relative to the total concentration of bicarbonate, carbonate, and hydroxide
(Gherini et al. 1984). These other acceptors in such instances include the following
species: organic compounds with carboxyl (-COOH) and phenolic hydroxyl groups (-0H), and
the monomeric aluminum species and their complexes.
Following a series of iterations developed by Tetra Tech Inc. (1984), ANC in
low alkalinity waters has been defined by the following equation:
ANC = Xcations - ^anions + 3A1-|- [21]
where,
Al^ is total dissolved Al (mol l"^) [22]
The above equation maintains the charge balance between cations and anions in solution.
In neutral waters, the concentration of dissolved aluminum is near zero which simplifies
the equation as: cations minus anions. If the ANC value is negative, this is indicative
of acidification. In acidic waters where dissolved aluminum is present, it must be
included in the calculations or the ANC value will be incorrect (Tetra Tech Inc. 1984).
Recently, Herczeg et al. (1985) developed a new method for the monitoring of
temporal trends in acidity which they claim avoids the problems of lack of sensitivity
common with current technology. Their method is based on calculations of the equilibrium
relationship between dissolved inorganic carbon and the partial pressure of CO2.
Herczeg et al . (1985) claim that this new method eliminates biases in pH as measured by
electrode, and minimizes the effects of natural perturbations in acidity caused by
biological activity and its effect on PCO2.
170
8.2 SENSITIVE WATERS
Hendrey et a1. (1980a) defined sensitive waters as those with alkal inities
below 200 yeq L ^, a level low enough to be neutralized by acidic deposition and runoff.
Gibson et al. (1983) refined this classification to give the following ratings:
less than 50 yeq L
50 - 100 yeq L^^
100 - 200 yeq
greater than 200 yeq L
- extremely sensitive
- very sensitive
- sensitive
- not sensitive
However, in other investigations, Canfield (1983) classified waters with alkal inities
between 200 and 400 yeq L ^ as moderately sensitive. Various researchers have predicted
the vulnerability of lakes and streams to acidification based upon these criteria (Gibson
et al. 1983; Haines et al. 1983; Scruton 1983; Kling and Grant 1984; and National
Wildlife Federation 1984) .
The calcite saturation index, CSI, has been suggested as an improvement over
alkalinity as a measure of aquatic system susceptibility to acidification (Kramer 1976;
Galloway et al. 1978). The general form of this index is represented by the following
formula :
(Ca) (HCOa)
CSI = -log [23]
(H)K
where,
() are ion molar activities and
K is the equilibrium constant CaCOa + H"^ = Ca^"^ + HCOa" [24]
Restructuring the formula gives the following form:
CSI - logK -log[Ca] - logCHCOa] - pH [25]
where log K = 2.582 - 0.242t, t being the temperature (°C).
Haines et al . (1983) have claimed that CSI is not superior to pH and alkalinity measures
and is needlessly mathematically complex. Since the derivation of CSI is calculated
from both pH and alkalinity, this seems to be a valid argument. Alkalinity is easily
measured and even historical values can be readily recalculated if necessary, and it
would appear that of the three types of sensitivity measurements available this should
be the preferred method.
8.3 WATERSHED CHARACTERISTICS DETERMINING SURFACE WATER SUSCEPTIBILITY TO ACIDIFI-
CATION
Lake acidification studies in the Adirondack Mountains, New York State, have
clearly shown that all surface waters are not equally susceptible to acidification. A
lake's vulnerability to atmospheric deposition depends upon its biogeochemi stry and
hydrology of its entire catchment area, including the type and condition of vegetation
cover, bedrock characteristics, and the type and depth of soil (Goldstein et al. 1984a).
171
A summary of the watershed characteristics that influence surface water susceptibility
to acidification and the precipitation pathways that must be considered when making an
assessment of sensitivity are shown in Table 34 and Figure 8.
8.3.1 Major Determining Factors of Surface Water Acidity
8.3.1.1 Forest Canopy. The forest canopy interacts with and changes the chemistry of
intercepted acidic deposition. This topic is dealt with at some length in Section 2 of
this report (Forest Effects) and is only mentioned here to reiterate the need to consider
this factor when assessing the impact or potential for impact of acidic deposition on
surface waters.
8.3.1.2 Bedrock Geology. Bedrock geology has been most frequently used to assess
potential susceptibility of surface waters to acidic inputs (Hendrey et al. 1980a;
Kaplan et al. 1981). The chemical characteristics of the bedrock generally determine a
region's susceptibility to acidification. Hendrey et al. (1980a) described four types
of bedrock which distinguish susceptibility to acidification.
Type 1. Granite/syenite, granitic gneisses, quartz sandstones, or metamorphic
equivalents. Low to no buffering capacity. High sensitivity.
Type 2. Sandstones, shales, conglomerates, high-grade metamorphic felsite to
intermediate igneous rocks, calcsilicate gneisses (no free carbon-
ates). Medium/low buffering capacity. Medium sensitivity.
Type 3. Slightly calcareous, low grade, intermediate to mafic volcanic, ultra
mafic and glassy volcanic rocks. Medium/high buffering capacity.
Low sensitivity.
Type 4. Highly f ossi 1 i f erous sediments or metamorphic equivalents. Limestones
or dolostones. High buffering capacity. Very low sensitivity.
Areas underlaid by limestone or other bedrocks high in calcite or other carbon-
aceous materials have extremely high buffering capacities capable of neutralizing
extensive loadings of acids. Conversely, areas underlaid by granitic or related igneous
rocks or their non-calcareous materials have an extremely limited buffering ability.
Norton (1980, cited in Marcus et al . 1983) states that even small amounts of calcareous
material in a watershed can exert enough buffering to change susceptibility from high to
moderate in some watersheds.
8.3.1.3 Soil Type and Depth. Except when it falls directly on to surface waters or on
to exposed bedrock, most precipitation percolates through or over soils prior to entry
into the receptor waters. During the course of this passage, the chemistry of the
waters can be changed substantially and this is the topic of the following section of
this report (Section 5: Soils). Direct precipitation, surface runoff, and lateral flows
from soil strata have their characteristic ranges of pH ranges and chemical species, and
these are shown in a general way in Figures 9 and 10.
172
Table 34. Watershed characteristics that influence surface water
susceptibility to acidification.
Category
i nc rea s eu
Susceptibi 1 ity
uec rea sea
Susceptibi 1 ity
Bedrock geology
Soils
Buffering capacity
Resistant to weather-
"inn ( mci+ami^K^nhn^
1 iiy ^ iiic tdiiio 1 pri 1 L y
igneous)
Low
Easily weathered (sedi-
niciiLdry, caicixe
containing)
High
Depth
Sha 1 1 ow
Deep
SO4 adsorption
capacity
Low
High
1 opog rapny
oxeep s lopea
oua 1 1 OW 5 lOpeu
Ratio of watershed
to surface water area
Low
High
Lake flushing rate
High
Low
Watershed vegetation
and land use
Dominant vegetation
Forest management
Con i f erous
Reforestation
Dec i duous
Clearcutting
Water quality
Alkal inity
1 ropn 1 c s xaxus
Cultural eutro-
phication
Low (<200 yeq/L-M
Highly oligotrophic
Forestry
High (>200 yeq/L"!)
Less oiigoLropnic,
mesotrophic, eutrophic
Agriculture, municipal
Humic substances
Absent
Present
Sphagnum moss
Present
Absent
Sulfate reduction
potential
Low
High
Climate/meteorology
Precipitation
Snow accumulation
Growing season
Alkaline dusts
High
High
Short
Low
Low
Low
Long
High
Source: Marcus et al . (1983)
173
174
Figure 9. Lateral flow of water from different soil layers in
determining lake water pH (after Chen et al . 1984).
Atmospheric
deposition
Organic
horizon
Thin mineral
horizon ^RCOOH
Drainage flow patti
H+,S042-,N03-
Surface water
Thick mineral
horizon AKOHlj |'
RCOO
NO,
Ca^+.SQ
4 1
NOj-jHCO:
Figure 10. Chemical species associated with water flow paths to
a lai<e (after Driscoll and Newton 1985).
175
8.3.1.4 Topography and Watershed-to-Lake Ratio. Both topography and watershed-to-lake
ratio affect the susceptibility of surface waters to acidification (Panel on Lake
Acidification 1984). During episodic events, these factors can influence or cause pH
depressions of lakes and streams.
Watersheds with steep slopes or with little vegetation cover exhibit rapid
runoff and are termed hydrological ly flashy. Because of the rapidity of throughflow, it
has been suggested that surface waters in such basins receive atmospheric precipitation
largely unchanged in its chemical composition (Marcus et al. 1983). In systems with
organic substrates through which the runoff passes either as overland flow or as perco-
lation causing piston flow, the previous generalization may not be true. In either
case, however, lakes or streams in such areas become more susceptible to acidification
provided acidic deposition is occurring. A smaller watershed-to-lake ratio also increases
susceptibility of lakes to acidification due to direct input of atmospheric deposition
on its surface without the intervention of modifying agents. This process has been
observed in Southern Ontario by Dillon et al. (1978).
Lakes in drainage areas with high watershed-to-lake area ratios tend to be less
susceptible to acidification. This is primarily due to the basin's attenuation capacity
and the longer time of contact with vegetation, soils, etc., all of which can ameliorate
the effects of incoming acidic precipitation.
8.3.1.5 Watershed Vegetation and Land Use. Rosenqvist (1978b) and Krug and Frink
(1983a, b) have claimed that terrestrial vegetation in watersheds markedly influences
surface water pH. The effects of vegetation on both dry and wet acidic deposition is
the topic of Sections 2 and 3 (Forests and Agriculture) of this report and should be
consulted by the reader for the types of mechanisms and processes that incoming acidic
deposition is subjected to, upon interception by vegetation and how this can have
significant changes upon its ultimate chemical composition as it enters receptor waters.
8.3.1.6 Surface Water Quality. Alkalinity is the most important factor that determines
the susceptibility of a water body to acidification. Waters with alkalinities lower than
200 yeq L ^ have high susceptibility to acidification while alkalinities greater
than 200 yeq L ^ provide their basins with low susceptibility to acidification
because of their high buffering capacity (Hendrey et al. 1980a). As discussed earlier,
bedrock geology, soil, changing land use practices, and vegetation types generally
influence surface water alkalinities. Runoff from agricultural lands which is high in
nutrients and often lime, or from clearcut forest stands, appears to increase surface
water alkalinity and pH, thus decreasing susceptibility to acidification (Rosenqvist
1978b; Braekke 1981). Decrease in alkalinity and pH can result from drainage originating
from reforested lands or from agricultural areas subjected to ammonium and sulphate-
containing fertilizers (Rosenqvist 1978b; Braekke 1981; and Hunt and Boyd 1981). In the
first instance, the causes for decrease in the pH are related to organic acid formation
by conifers and in the second, a result of microbial sulphur oxidation to sulphuric
acid. These topics are dealt with further in Sections 2 and 6 of this report.
176
8.3.1.7 Climate and Meteorological Conditions. High precipitation areas are generally
more susceptible to acidification because of the high leaching rate and consequently low
cation exchange capacity usually found in such regions. This allows acidic input to
pass through the system into receptor waters fairly quickly. Episodic events also
contribute to acidification, in particular, snowmelt is considered important. Winter
accumulations of acidic deposition and naturally produced acids are released to surface
waters in fairly short pulses during snowmelt and have been observed to produce acid
shock as a result, particularly in fish. Alkaline dust in the atmosphere can also
influence the acid-base status of lakes. Increases in the quantities of alkaline dust
tend to reduce the acidity of atmospheric deposition, a frequent phenomenon throughout
the western portion of North America (Marcus et al . 1983).
8.4 ACIDIC WATERS AND THEIR REACTION PRODUCTS
Acidification of waters may be defined as a loss or decrease in acid neutraliz-
ing ability as measured by alkalinity. The major source of alkalinity in water is the
bicarbonate ion which is produced by carbonic acid weathering of the surrounding bedrock
and soil. The most common reactions resulting in weathering of limestone, dolomite, and
silicates are:
1 . H2O f CO2 ^ H2CO3 [26]
2. CaCOa -I- H2C03 -> Ca(HC03)2 or Ca2+ + ZHCOa" [27]
3. CaMg(C03)2 + 2H2CO3 Ca^^ + Mg2+ + 4HCO3- [28]
4. CaAl2Si 208 + 3H2O + 2CO2 Ca^"^ + 2HC0^" + Al 2Si 2O5 ( OH) [29]
In recent years, alkaline lakes considered to be sensitive, however, have been
reported as acidic due to atmospheric deposition particularly of sulphuric and nitric
acids.
Acidic substances in the atmosphere can reach surface waters through three
pathways: (a) direct deposition from the atmosphere as dry and wetfall; (b) indirectly
via runoff over or through the watershed; and (c) through internal generation within the
watershed itself, for example, acidification of soils which in turn leads to the
acidification of the water (Marcus et al . 1983; and Mason and Seip 1985). If, as a
result of the loading of H^, the overall alkalinity is depressed, the mobilization of
potentially toxic heavy metals may also occur. For example, if pH is depressed to
values between 4 and 5, ionic forms of aluminum, zinc, and lead, to name a few, may be
released from the sediments. Above a pH of 6.5, these metals generally precipitate out
of the water and are also absorbed by the sediments.
Acids deposited on the land result in competition between H^ and COs .
The net effect of this action is the leaching of bicarbonate, calcium, and magnesium
ions causing the soils to lose buffering ability and eventually reducing the pH of the
soil. As the atmospheric deposition continues and mineral, strong acid-weathering takes
over from carbonic acid weathering, metals such as aluminum become mobilized. In this
instance, the products reaching the aquatic environment are not H^ ions but byproducts
of reactions caused by soil acidification. Other types of alterations may also occur.
177
most notably in the cation exchange processes in the soil, whereby H is exchanged
with metal cations. It has been suggested that cation exchange occurs very rapidly if
the H"^ is in solution, whereas the release of cations from other mineral sources is
slow (Panel on Lake Acidification 1984). These types of reactions involving cations
occur only in soils in which the cation exchange capacity is higher than 20 yeq/100 g ^.
Below this value, h"*" may be exchanged more slowly or not at all, thereby accumulating
in the soil. If the concentration continues to increase and the soil pH drops
below 6.0, toxic metals such as aluminum become more soluble and are mobilized.
Aluminum, for example, may be mobilized as free aluminum, or as a complex with fluoride,
hydroxide, sulphate, or organic ligands (Driscoll 1980). The solubility of aluminum and
other metals is highly pH dependent and its inorganic form determines whether its
concentration increases with either increasing acidity or alkalinity (Driscoll 1980).
Hydrogen ions may also be generated internally by humic substances in the soil.
This internally produced h"^ can also mobilize heavy metals by means similar to those
outlined above for weathering (Krug and Frink 1983b).
The contribution of to the aquatic environment is, therefore, the sum of
internally produced ions plus those remaining after weathering and ion exchange pro-
cesses. This portion of acidity is combined with leached metal ions and mobile anions
such as sulphate, nitrate, and organic anions. Since, in most cases, only a small
fraction of acidic deposition falls directly on a lake relative to the land surface,
surface water acidification depends primarily (but not solely) on the flow path that
precipitation follows in a watershed prior to reaching the water body.
Cation and anion balance is the rule when ions are transported from soil to the
aquatic environment so that electrochemical neutrality is maintained. Anions, such as
sulphate and nitrate, are provided by acidic deposition. Although sulphate ions
(S04^ ) are quite mobile, their mobility in a soil depends on its sulphate adsorp-
tion capacity (SAC). Most sulphate will be retained if the SAC is undersaturated
(Galloway et al. 1984). Drablos and Tollan (1980) have reported that little retention
of sulphate occurs in many North American or Scandinavian watersheds. Nitrate ions
(NO3 ) are also quite mobile, but their mobility depends on biogeochemical processes
such as uptake by vegetation (Overrein et al. 1980) and on the velocity of water
movement. During storm events when discharge and velocity of water increases, nitrate
ions become highly mobile. However, in most instances, sulphate tends to be the major
anion balancing cations with contributions being made by organic acid anions from
internal generation or via acidic deposition. By means of oxidation processes, naturally
occurring minerals such as pyrite, and NH4"^, can also contribute sulphate and
nitrate ions.
In summary, in areas receiving neither acidic deposition nor marine salts in
precipitation, calcium and magnesium should be derived solely from weathering and leac-
hing processes associated with carbonic acid and organic acids. The addition of calcium
and magnesium ions in such cases should be electrochemical ly equivalent to alkalinity.
In areas receiving acidic deposition, additional calcium and magnesium may be leached a-
nd an excess of hydrogen ions may cause reductions in alkalinity. Lakes in such water-
sheds exhibit calcium plus magnesium equivalences greater than alkalinity, and the
excess cations are balanced by non-marine sulphate and nitrate (Aimer et al. 1974; and
Dickson 1980).
178
8.5 PRECIPITATION QUANTITY AND QUALITY AS FACTORS IN SURFACE WATER ACIDIFICATION
As stated previously, there are three general pathways for acidic substances to
enter aquatic systems. Disagreement exists in the literature due to the lack of quanti-
tative data documenting the contribution from each source. Gorham and McFee (1980)
recognized this deficiency and suggested mass balance studies to determine the origins
of acids, metals, and organic molecules entering aquatic ecosystems. Several studies
are attempting to do this either by direct measurement or by predictive estimation using
models (Dillon et al. 1982; Wright 1983; and Gherini et al. 1984). Of these studies,
Gherini et al. (1984) using the model developed in the Integrated Lake-Watershed
Acidification Study (ILWAS) actually attempted to quantify contributions for each routing
in a mass balance.
In ILWAS a model was developed for two lakes in the Adirondack Mountains.
Simulation of the model showed that routing of waters through soils (shallow versus deep
flow) largely determined the extent of lake acidification. Analysis of the basins of
the two modelled lakes, combined with field data, indicated that the internal production
of acidity was approximately two-thirds the amount of atmospheric loading.
Likens et al . ( 1 977 ) observed that hydrogen ion concentrations in Hubbard Brook
were directly correlated with discharge volumes (r^=0.73). They further suggested
that observed increases in could be partially explained by the amount of precipita-
tion. Rosenqvist (1978b), working in southern Norway, reported that following a storm
event the pH in a stream dropped from 5.6 to 4.4, resulting in a pH shift equivalent to
5 times the acidity of rainwater entering the system. He attributed the pH shift to
cation exchange processes in the soil and increases in leaching rates of naturally
occurring hydrogen ions due to overland runoff and shallow groundwater throughflow.
The observations of Likens et al. (1977) and Rosenqvist (1978b) suggest that
changes in stream chemistry during and immediately following precipitation events are
greatly influenced by soil chemistry rather than solely by the H^ concentration in the
incoming rainfall. For example, large amounts of H^ and Ala^ occur in soluble
form in the naturally acidic environment of a coniferous forest, and all that is required
to flush these chemicals through the system is a storm event of sufficient magnitude.
Large amounts of these materials will be washed through the 0 and A soil horizons, where
they are mostly produced, particularly if the terrain is steep and the storm is intense
enough .
According to Elzerman (1983), changes in the chemical composition of stream
water during precipitation events are important for the following reasons:
1. Significant portions of total annual fluxes of dissolved and particulate
materials can occur during and following precipitation events;
2. Some products of neutralization and other reactions of rainwater with
watershed components will appear in stream waters; and
3. Episodes in chemical composition above threshold values may occur; for
example, transient events in which aluminum reaches toxic concentrations.
179
Elzerman (1983) studied the effects of precipitation events on the chemical
regime of a watershed in South Carolina. He observed that during the precipitation
events, in comparison to pre-event levels, pH dropped slightly (0.6 units), as did the
concentration of bicarbonate ion (198 to 147 yeq L ^). The major cations, Mg^^, Ca^^, and
Na^ were reduced in concentration as a result of dilution and this in turn accounted
for the slight acidification of the stream. The concentrations of sulphate and aluminum,
on the other hand, increased from 8.5 to 143 and 10 to 100 yeq L ^, respectively.
The quality of incoming precipitation can also potentially affect surface water
chemistry. Lunde et al. (1977) detected more than 450 organic compounds in precipitation
over Norway. These compounds, all of which were suspected to be from anthropogenic
sources, consisted of: alkanes, polycyclic aromatic hydrocarbons, phthlates, fatty
esters, aldehydes, amines, pesticides, and polychlorinated biphenyls (Lunde et al. 1977;
Strachan and Huneault 1979; and Alfheim et al. 1980). Haines (1981) suggested that they
were likely involved in acidic precipitation. However, such compounds, unless present
as organic acids, do not contribute to surface water acidification.
Within Canada, impacts of atmospheric deposition on watersheds are most apparent
in association with point source emissions. The best documented cases are found in the
area surrounding Sudbury, Ontario (in relation to smelter operations) and in Halifax
County, Nova Scotia. Harvey (unpublished work in Beamish 1976) found that sulphate
concentrations in over 100 lakes in the Sudbury area were indirectly related to the
distance from sources. Nickel and copper concentrations in lakes of the La Cloche
Mountains, southwest of Sudbury, were found to exhibit similar trends (Beamish 1976);
ranges in nickel and copper concentrations were 5 to 15 and 2 to 4 yeq L ^, respectively,
in affected lakes and less than 3 and 2 yeq L ^, respectively, in remote lakes.
Watt et al. (1979) compared concentrations of hydrogen and sulphate ions in
116 lakes from Halifax Country, Nova Scotia and found results similar to those reported
by Beamish (1976), that concentrations decreased with increasing distance from emission
sources. In comparison to studies conducted 21 years previously, Watt et al. (1979)
also found that the lakes were significantly more acidic. The decrease was related to
the geomorphology of the lake basins studied. They found that pH decreased by approxi-
mately 0.34 units in granitic basins and by 0.65 units when the basin was composed of
metamorphic rock. The 1955 pH values of the lakes studied indicated that none were well
buffered and had original mean pH values' of 4.66 on granitic and 5.62 on metamorphic
rock. Similar results were obtained for sulphate concentrations. In granitic basins,
sulphate concentrations increased by about 27.91 yeq L ^, whereas in metamorphic basins,
the increase was 47.5 yeq L ^.
In addition to the aforementioned ions, increased concentrations of heavy
metals have also been identified in relation to the loadings of point source emissions.
Metals most often affected are lead, zinc, manganese, iron, nickel, mercury, vanadium,
aluminum, and cadmium (Gorham and McFee 1980). Haines (1981) observed that metal con-
centrations were higher under conditions of acidic rather than non-acidic precipitation.
Metal concentrations close to emission sources were the highest, but elevated concentra-
tions were also observed far from identifiable sources. This suggests either long-range
transport or increasing scavenging from the atmosphere as a result of acidic precipita-
tion, or due to dry deposition or a combination of all three. Tomlinson et al. (1980)
180
pointed out that increased acidity of precipitation strips mercury from the atmosphere,
the mercury likely originating from natural sources.
Snowmelt also affects surface water quality and has been studied by numerous
authors in the context of acidic deposition (Galloway et al . 1980; Hendrey et al . 1980b;
Johannes and Altwicker 1980; Johannessen et al . 1980; Bjarnborg 1983; Cadle et al. 1984;
and Schofield 1984). Snow pH may be as low as 3.3, but in general values range from 4.5
to 5.0 (Seip 1980). The rapid release of acids from snow during a thaw can cause a
rapid drop in the pH of poorly buffered lakes and streams. This phenomenon is termed
acid shock and can have severe and drastic effects on aquatic life (Schofield 1976a, b;
Hultberg 1977). Studies have shown that sulphate is preferentially leached from snow-
packs during the winter making nitrate the dominant ion during the melt (Johannessen
et al. 1980). Nitrate is biologically active. However, it is difficult to ascertain its
importance in stream and lake acidification. Aging snow has also been found to become
progressively less acidic, perhaps as a result of cation exchange with organic debris
entrapped in the pack (Hornbeck et al. 1977).
During the first stages of the spring flush, higher concentrations of ions have
been detected in stream waters in comparison to the snow pack itself (Hendrey et al.
1980b; Johannes and Altwicker 1980; and Cadle et al. 1984). As the snow pack melts,
significant proportions of the major acidifying species are released during the loss of
the first 21-35% of the pack. It has been suggested that melt occurs in three stages
which can explain the resulting effects on stream water chemistry (Johannessen et al.
1980). During the first stage, old basin waters are pressed out of the system by piston
flow which releases weathered ions such as calcium, magnesium, and bicarbonate, causing
their concentrations to rise. It is followed by dilution of the ions during the second
stage by the rising meltwaters containing low concentrations of weathered ions but high
concentrations of acidic ions. The third stage consists of the remaining snowmelt which
is relatively dilute in ionic content, thus causing a further dilution effect.
Jeffries et al. (1979) showed that hydrogen ion discharges from Canadian
watersheds varied proportionately with discharge volumes during the two-month snowmelt
period. It was found that runoff acidity was not relatively high in the early periods
of snowmelt as was expected. This suggests that sources in addition to the snow pack
accumulation, such as soil leaching, contributed to the discharge hydrogen ion loadings.
Seip (1980) also felt that increased hydrogen ion concentrations in snowmelt waters
reflect not only snow accumulations but also the influences of accumulations over winter
in the soil. Soi 1-meltwater contact is extremely important in determining the final
chemistry of thaw waters. Factors identified by Seip (1980) as contributing to this are
soil freezing, air temperature, thickness of snow cover, texture of soil, and type of
vegetation. Vegetation is most important in organically rich areas or in areas such as
coniferous forests where the litter layer is naturally acidic. Seip (1980) states that
this type of situation results in the leaching of organic acids which can then contribute
substantial quantities of hydrogen ion to the meltwater.
Schofield (1984) reported on the temporal acidification of three lakes in the
Adirondack Mountains during snowmelt. He attributed the observed effect to base cation
dilution and to increased strong acid associated anion levels, particularly nitrate.
181
These changes in surface water chemistry were related to an upward shift in flow paths
from groundwater dominated base flow (mineral horizon) to shallow humus layer flow
during increased snowmelt as the shallow layers became saturated.
The literature surveyed clearly indicates that episodic events and factors of
precipitation quantity and quality affect surface water acidification. The literature
is still not definitive on how actual acidification occurs or the processes by which
surface waters are acidified.
8.6 POTENTIAL SOURCES OF ACIDIFICATION OF SURFACE WATERS
Considerable controversy exists in the literature regarding surface water
acidification. Most of the discussion revolves around the cause and effect relationships
of acidification with regard to origin of the problem and whether it is anthropogenic or
natural. Research that supports the theory that acidification is mostly anthropogeni -
cally derived may be found in Beamish et al. (1975), Overrein et al. (1980), Rahel and
Magnuson (1983), and Somers and Harvey (1984). The theory that land use practices and/or
internal proton production in soils cause acidification is supported by Rosenqvist
(1978a, b), and Krug and Frink (1983a, b). Following a review of both types of literature,
we feel that both points of view are valid but that anthropogeni cal ly induced land-soil-
water acidification likely accelerates natural acidification processes in some instances.
8.7 TRENDS IN SURFACE WATER ACIDIFICATION IN NORTH AMERICA
Numerous studies conducted throughout North America were reviewed by Telang
(1987). Only a few of these will be discussed in this overview and the reader is
referred to the main body of Telang's text for further information on the studies
outlined in Table 35.
In examining the trends in surface water acidification one must keep in mind
the following points when comparing data sets:
1. Limited data are available for poorly buffered systems (Marcus et al.
1983);
2. Prior to 1955, pH was measured using colourimetric methods; more recently,
potentiometric methods have been used;
3. Alkalinity was rarely measured prior to 1955 (Haines 1981);
4. Temporal differences and the effects of biological processes on pH were
often ignored in historical and predictive studies when comparing data
sets. For example, photosynthetic rates which reduce surface water
acidity, and respiration which causes an increase in pH were factors often
ignored. Studies conducted in Vermont on a soft water lake by Allen
(1972) illustrated the importance of such factors. He found that pH
changed from 5.65 to 9.57 between 0800 and 1200 hours and then decreased
to 6.35 by 1600 hours on the same day.
5. Changes in land use practices were often ignored when comparing data sets
and changes in acidity.
182
Table 35. Surface water acidification studies reviewed by Telang
(1987). ^
Ontario (Beamish 1974, 1976; Somers and Harvey 1984; Keller
et al. 1980; Glass et al. 1981)
Quebec (Jones et al. 1980: LaChance et al. 1985)
Atlantic Provinces (Watt et al. 1979; Scruton 1983)
Saskatchewan (Liaw 1982)
New England States (Haines et al . 1983; Marcus et al . 1983;
Davis et al. 1978; Hendry et al . 1980a; Norton et al . 1981;
Likens et al . 1977).
The Adirondack Mountains (Schofield 1976a, b; Pfeiffer and
Festa 1980; ILWAS (Tetra Tech 1983, 1984; Driscoll and Newton
1985)
New Jersey (Johnson 1979a, b; Morgan 1984)
Virginia (Shaffer and Galloway 1982)
South Carolina (Elzerman 1983)
Florida (Brezonik etal. 1980; Canfield 1983; Flora and
Rosendahl 1982)
California Sierra Nevada (Tonnessen and Harte 1982; Melack
et al. 1982)
Upper Great Lake States (Eilers etal. 1983; Rahel and
Magnuson 1983; Glass 1984)
Colorado Rocky Mountains (Lewis 1982; Turk and Adams 1983;
Gibson et al. 1983; Harte et al. 1985; Kling and Grant 1984)
183
6. Most studies, except ILWAS (Tetra Tech Inc. 1983-84) have not included
investigations of the major flowpaths that incoming precipitation follows
within a basin prior to reaching a lake, or the major processes that take
place along these paths that can and do alter the chemical characteristics
of the throughflow waters.
7. The parameters of colour and dissolved organic carbon were often not
included in studies of the trends in acidification of lakes and streams.
The parameters most commonly used to assess lake acidification have been
acidity, alkalinity, concentration changes in base cations, and changes in aluminum
concentrations. Some investigators report a trend in surface water acidification by
comparing historical data and recent data on pH and loss of alkalinity, while others
report lakes that may be susceptible to acidification based on bedrock characteristics
and the present day lake or stream chemistry.
Surface water acidification studies in North America can be categorized into
three groups. The first group consists of lakes for which the cause of acidification
has been attributed to long range atmospheric transport and inputs of acidic substances.
Studies conducted in the Atlantic Provinces, New England States, and New York State fall
into this category. Within the second grouping of lakes, acidification has been
attributed to the internal generation and contribution of hydrogen ions. Studies in the
New Jersey Pine Barrens fall into this category. The third grouping consists of lakes
for which acidification has not been established but predicted based on water chemistry
and basin characteristics indicative of susceptibility. Lakes in the California Sierra
Nevada and the Colorado Rocky Mountains fall into this category.
Keller et al.(1980) studied 200 lakes within a 200 km radius of the Sudbury
smelter area and found that 30% of the lakes had a pH less than 5.5, while 40% had
calcite saturation indices between 3 and 5, which are indicative of high sensitivity to
acidification. The cause of acidification was attributed to smelter operations. Glass
et al. (1981) suggested that about 6% of 1527 lakes surveyed in Ontario may be classified
as acidic. Alkalinities in these acidified lakes were near or less than 0 yeq L~^. Of
the 103 acidified lakes, 83 are subjected to local loadings from smelter operations in
the Sudbury and Manitoulin areas of the province. In the La Cloche Mountains, near
Sudbury, Beamish (1974, 1976) found evidence of fish population losses which he attribu-
ted to acidification and heavy metal toxicity. He found that although concentrations of
base cations did not show a significant difference between acidified and non-acidified
lakes, calcium concentrations were twice as high in the first category of lakes.
Acidified lakes also showed an overabundance of sulphate which represented 90% of the
anion content of these lakes compared to values around 40% in non-acidified water
bodies. These studies clearly indicated that lake acidification was caused primarily by
local point source loadings from the smelter operations at Sudbury and the surrounding
area.
Recently, Driscoll and Newton (1985) sampled twenty lakes and their watersheds
in the Adirondacks in an effort to evaluate mechanisms that control the sensitivity of
lakes to acidification. Two of the lakes were of the seepage variety with no inflow or
outflow, and the other 18 were drainage lakes. The two seepage lakes and three of the
184
drainage lakes were then selected to illustrate the range of chemical composition found.
The two seepage lakes received most of their water directly from precipitation and
therefore were considered highly sensitive to acidic deposition. The pH values of the
lakes were 4.7 for Barnes Lake and 4.3 for Echo Lake. Because these lakes received no
runoff waters and were isolated from the groundwater system their concentrations of base
cations and dissolved silica were very low. Despite their low pH values, both lakes
also had low concentrations of aluminum. Differences in the physical and chemical
characteristics of the lake basins were reflected in their chemistry. Barnes Lake,
which is a perched clear water lake, received most of its acidity from atmospheric
deposition and resulting in high sulphate concentrations. Echo Lake, however, while
also receiving some acidic deposition, received the majority of its acidity from peat
deposits, up to 28 metres thick, that surrounded the lake and released organic acids
into its waters.
Of the three drainage lakes examined by Driscoll and Newton (1985), Merriam
Lake was acidic (pH ^.5) as a result of a lack of base cations in its catchment area;
West Pond was a bog lake (pH 5.2) whose acidity was mainly attributable to organic acid
drainage plus acidic inputs via the atmosphere; and Cascade Lake (pH 6.5) was relatively
insensitive to acidic deposition because of the balance maintained between atmospheric
sulphate additions and high concentrations of base cations in its waters.
As a result of these studies, Driscoll and Newton (1985) suggested that sensi-
tivity of lakes to acidification varies from lake to lake and that it depends on
hydrology, mineralogy, and vegetative cover in the study area. They suggested that in
the Adirondack lakes, organic acids were the main cause of brown water lake acidity, and
that sulphuric acid, and to a lesser degree nitric acid deposition were the main causes
of clear water acidity.
The longest continuous data base on stream water and precipitation chemistry in
North America relates to the Hubbard Brook Experimental Forest in New Hampshire. This
data base extends from 1964 to 1 974 and has been reported by Likens et al. ( 1 977).
During the 10-year period, Hubbard Brook received acidic deposition with an average
precipitation pH of 4.12. In addition, the area is classified as highly sensitive to
acidification due to its bedrock geology and soil characteristics which indicate
extremely low buffering ability. In spite of these factors. Likens et al. (1977)
reported the maintenance of relatively constant stream water chemistry and no apparent
trends in the stream pH, which remained near 5.0 over the study period.
8.8 EFFECTS OF ACIDIC DEPOSITION ON AQUATIC BIOTA
Acidification of freshwaters is a complex process involving not only increases
in acidity but also increases in metal ion concentration, increased water clarity,
accumulation of periphyton and detritus, changes in trophic interactions such as loss of
fish as predators, and physiological changes in aquatic organisms. Magnuson (1983) has
suggested that a study of the response of aquatic systems to acidic deposition must take
into account all these types of changes because together they constitute the acidifi-
cation process.
The impact of acidic deposition on aquatic biota was first observed in fish
populations with the earliest record being from Norway where Atlantic salmon populations
185
began to decline in the 1920's (Jensen and Snekvik 1972). Since then, declining fish
populations have been reported for many lakes and rivers throughout the world in areas
receiving acidic deposition (Beamish and Harvey 1972; Drablos and Tollan 1980; and
Schofield 1982). These fish extinctions were related to changes in water chemistry,
particularly increases in acidity and heavy metal concentrations. Later, studies on the
effects of acidification on aquatic organisms were broadened to include other trophic
classes of organisms as well as fish.
Telang (1987) has reviewed this literature by first identifying those organisms
commonly found in naturally occurring acidic waters, and then by documenting the experi-
mental evidence for acidification effects caused by deposition. Organisms that occur in
naturally acidic waters are shown in Table 36. This type of information provides a
baseline from which to measure the effects of natural versus anthropogenical ly caused
acidification. The results of numerous studies into the effects of acidic deposition on
aquatic organisms may be found in Table 37 which indicates that the effects are varied
and definitely not restricted to fish. In general, it appears on the basis of the
experimental evidence (Table 37) that effects of surface water acidification on aquatic
organisms becomes apparent as alkalinities decline to 65 yeq L ^ and pH values
decline to about 6.0. Baker (1983a, b) suggested that pH declines below this value would
result in escalating biological changes such as loss of benthic species from the
community, loss of fish species, and loss of community structural complexity. This last
point is important because it would indicate that the resilience and ability of the
biological community to withstand further change would also be lost such that future
impacts may cause an exponential effect on overall system productivity. The effects of
acidification also become apparent during episodic events when the pH of the water may
drop from as high as 7.0 to 4.3 very quickly. Such an episodic event in many areas of
North America results from snowmelt. Although acidic deposition can have detrimental
effects on aquatic biota, it is only one of the many factors working jointly in the
ecosystem to produce such effects. Others include natural acidic waters and acid
forming processes, elevated levels of toxic metals, availability of nutrients, and
changing land use practices.
8.9 MODELS OF FRESHWATER ACIDIFICATION
Three types of models have been developed to evaluate the acidification of
freshwaters by acidic deposition. These include empirical models, static and dynamic
mechanistic models, and conceptual models. Telang (1987) has reviewed a number of these
modelling efforts from each category and found problems in all cases, with the possible
exception of the ILWAS model. The ILWAS model is a dynamic model that is suggested to
have potential for universal applicability. Only the ILWAS model will be discussed in
the following section. The reader is referred to Telang (1987) for further details of
the other models reviewed.
8.9.1 ILWAS Model
The Integrated Lake-Watershed Acidification Study (ILWAS) model was developed
to provide a mathematical approach capable of predicting changes in surface water acidity
given changes in the acidity of precipitation and dry deposition (Gherini et al . 1984;
186
Table 36. Lower pH limits for various groups of organisms in naturally
acidic waters.
Approximate
Lower pH Examples of Organisms
Group Limit Occurring at Lower pH Limit Reference
Bacteria
Fungi
Eucaryotic algae
0.8 Thiobacillus thiooxidans , Brock (1978)
Sulfolobus acidocaldarius
2-3 Bacillus, Streptomyces Brock (1978)
0 Acontium velatum Brock (1978)
0 Cyanidiwn caldariwn Brock (1978)
1-2 Euglena mutabilis, Brock (1978)
Chlamydomonas acidophila,
Chlorella
Blue-green algae
Vascular plants
Mosses
Protozoa
Roti f ers
Cladocera
Copepods
Insects
Amphipods
Clams
Snai 1 s
4.0
2.5-3
3.0
2.0
3.0
3.5
3.0
3.0
3.6
2.0
3.0
5.8
5.8
4.5
6.0
5.8
6.2
Mastigocladus, Synechococcus Brock (1978)
Eleocharis , Car ex,
Ericacean plants
Phragmi tes
Sphagnum
Amoebae, Heliozoans
Brachionus, Lecane, Bdelloid
Colotheca, Ptygura
Simocephalus , Chydorus
Macrocyclops
Cyc I ops
Ephydra thermophila
Chironomus riparius
Mayf 1 ies
Gammarus
Pisidium
Most other species
Amnicola
Most other species
Brock (1978)
Hargreaves
et al. (1975)
Ueno (1958)
Brock (1978)
Brock (1978)
Hutchinson
et al. (1978)
Edmondson (1944)
Ueno (1958)
Ueno (1958)
Hutchinson
et al. (1978)
Brock (1978)
Hutchinson
et al. (1978)
Sutcliffe and
Carrick (1973)
Sutcliffe and
Carrick (1973)
Griffiths (1973)
Pennak (1978)
Pennak (1978)
Pennak (1978)
continued .
187
Table 36 (Concluded) .
Approximate
Lower pH Examples of Organisms
Group
Limit
Occurring at Lower pH Limit
Reference
Fish
3.5
Tribolodon hakonensis
Mashiko et
al.
(1973)
4.0
Umbra limi
Rahel and
Magnuson
(1983)
4.5
Sunfishes (Centrachidae)
Rahel and
Magnuson
(1983)
Source: Magnuson (1983).
188
+->
c
O)
E
(O i-
_i OJ
Q.
X
CO I —
o cn
-a Q.
c =J
o o
o t_
o
•I— «/1
-o •<-
•r-
cn
J- c
<U -r-
-O -1<£
C 03
=3 E
3 3=
Q) Q.
U_ I—
. o s-
31 ■— a;
Q. Q) >
^ <u
o ■«->
r- c i-
<D o
O M-
+J CD
s- +->
<U Q. Q.
> <U
•r- 1/1 <_J
QJ X
•1- .r- O
CD O
CL •
— 't—
I— l/»
ftl i- •
c q; o
i- Q. .
<D in
cn > I
c: tn
r- E •
O Q.
x:
i/» 3 o
QJ +->
I— to
<-> E c
O) fO 3
Q.r— O
C>0 O "O
I- -M tn
C
1
1— 1
o
u>
X
1/1
CJ>
a;
o
•o
c:
(/T
o
cz
fO
(1)
-a
c:
03
+->
o
i_
c
o
>
na
=3
in
fa
i-
o
□c
+->
"O
3:
<u
o.
fO
CL >, 3:
E
i_
i_
o
CD
ro
i_
CI
o
3
o
1/1
+->
o
<U
4^
o
+->
rO
r—
03
=3
QJ
i-
' ^
c:
X3
o
03
E
o
CO
pi
Z3
+->
i_
+->
(0
C\J
1/1
Q
<D
(—»
Ln
C
l/l
U
rO
<U
<D
f—
o
pi
1/1
<D
<U
(0
c^
x:
3:
Q
i_
o
C/1
<t
+-»
Q.
fO
r~
i_
cr»
o
f—
CT>
lO
cn
E
>i
O)
P~-
O
a>
cn
S-
1 —
>>
+->
-a
Qi
fO
l/l
c
+->
01
O)
1/1
i-
■JO
o
o
O)
s:
C_3
CO
to
cr
to
c
o
>>
CO
r— to
O) >
•r- i-
Ll_ CU
O
to
CU CD
J- cn
CU f—
E ^
O r-
CXD oa w
r— <U JxJ
T3 -r- i-
C .— 1-
tO o to
!—+->(_)
O 00
•> to
1/1 "O
E •«-
H +-> 03 -r-
CLf— i-
o a> o CU
E o ra
X f— XI
CU •!- Q.
1/1 to (/)
i- >»
<U r—
Ol-r-
r— CU C E
O CL-r- 03
5; (/) <4_ i4_
(J l/l
1/1 CU
=3 T-
>)
to
s_
C_)
1/1 1/1
•O =3
O $-
Q. to
•r- E
^ E
O. 03
E CD
<: ^
to
<U i-
03 CU
> +->
S- Q.
03 O
r— S-
CU
>> E
r— CU
M- ^
>> cn.
03 UJ
189
"O
(Z
0.)
CJ
*^
"O
**"" ^
c
fO
4-*
+-' o
<D
o
fO
•
■
•1 —
T3
(_)
13 +->
>>
d
<l>
k-
00
E
*'
O
D" C
-H
fO
(U
Q.
•r—
O
"O
•r—
' '
JC
Z3
fO
00
^
o
QJ
00
<4_ jQ
•r-
1/1
O
i_
CD
00
O 03
<u
CD
C
C
OJ
fO
fO
s_
a>
•t —
c
00 O)
<u
<u
00
x:
Q. i-
M-
c
+->
+->
XI
03
u
Z3 03
O
M-
fO
CU
O
■o
03
s_
cu
S_
$- 00
-M
a>
O)
XI
>
u
CJ) E
C
00
<u
ro
c
00
00
CO
fO
u-
O
+->
x:
•1—
QJ
<1> •!-
s_
Oi
s_
S- C
O)
i-
•o
in
x:
LJ
•r- 03
o
•r —
>
00
+->
CD
+-> CD
o
c
O
n:
•r~
Q.
c s-
H-
CO
Q.
3
00
UJ O
CO
■o
1
(D
00
<u
•
-S
>>
sz
<U
•♦->
•
o
fO
1
00
o
E -o
•
00
E
Z3
+->
o
o
<1>
a>
<_)
s_
■M
•r—
CU
XJ
00
•1—
(U
+->
M-
13
x:
•H
4-
X3
O
•r-
c
■o
00
X
i_
o
o x»
-o o
<D
•1 —
03
03
00
s-
•r—
+->
+->
O
M-
-H
O.
00
CD
J-
in
c
c
E
>>
o
i_
<u
03
o
(U
03
i_
O
CD
>
OI
o
00
S_
■H
o
00
>
o
C
Ol
O)
u
d
03
CD
o
o
•1—
s_
cn
■H
03
i-
=3
Q.
O
C
cu
a>
XI
00
■o
00
o
<U
<u
00
•r—
+->
t_
+->
c:
CD
J_
■o
E
a.
o;
03
03
+->
CD
o
O
03
=3
E
03
X3
>>o
Q.
■o
l_
00
i_
•f—
00
>»
+->
+->
-M
00
S-
in
sz
Q}
O)
CD
03
c
00
<D
o
Q.
•r —
E
J-
■a
CD
03
X3
o
IC
E
o
ro
(U
cr
03
o.
Q.
03
M-
<_>
ro
<C
Q.
00
1 —
XJ r-
z en
O0 C
o
CD
CD O
r— 00
^ CD
03
i-
<D
+-> r—
CD -a
c
f— x:
03 o
re (/^
^ CD 00
CD 03 CD
03 XJ T-
•O •> 1—
•I- X CD M-
+->•!- 03 C
1 u I- -o O
>>
^ CD O T- CD
+->
CJ C CJ <_) 03
00 03 O - — 00 i_
c=
t- XI +-> •«- -o
a>
CD O C +->
fO ' — »
■O « Z CD >> «
CD C=
•r- ^ E O ^ •
O
r— O
S_ <D ■»-> CD
CJ
CD
03 +->
+-* (U i/i ro fO
l_
>>
to "O i- O 00 X3 -M
00
U SI
•r- CD X5 CD •!— ro
o
•r- Q.
»- I- E J- r- C C
x:
O
^ •!-
CD i- E CD ■»-> 03
+->
+-> i-
+J a> •!- -M CD t- T3
c
i_
C CD
ro CD 3 03 CD >> O
CD
+->
CD O.
3 ^ «/l 3 XJ CD
CTJ
00
CO >— '
190
(/I
•
0)
• ft
•
<u
00
+->
>^
•r—
to
-»->
irj
00
•
4-»
c
1/1
0)
r-
•
1
T3
s_
• ft
13
•
•r-
o
<TJ
«/)
in
-o
00
•1—
JZ
C
C
c
•r-
<U
O
•r—
■M •
<U
o
•>
-M
o
3
(U
-H
S_
J-
E
U
C ID
•r- >>
ro
S-
c
+->
CU
E
o
Z3
to
o
o
fO
OJ •
U +-»
CD
1 —
•(—
Q
E
S-
l/l
S-
<D
u-
00
i-
00 Ln
Q) -r-
-a
r—
>
E
fO
C
o
cn
00
+->
•a
o
<U
«4-
Q.T3
fO
o
Z3
o
a;
00
•r-
•p-
+^
j_
+^
'
O
1—
0)
ZIZ
00
TCJ
OO
<_J
na
o
O
fO
n3
Q_
CD
c
(U
+->
>
u
a.
•r-
<U
<U
*+- fO
I—
c:
$_
i_
o
M-
c
O
c
00
s_
£=
00
C O
O
■H
CU
CU
00
o
H-
fO
o
t/i
O)
o
•r-
CD
(0 '
00
00
d
fO
00
■o
UJ
1 —
(U
I/)
o
fO
sz
(U
•r—
•r—
00
CU
d
O
Ql
E
+->
+->
■o
U
U
CU
00
i_
Ol
+->
■a
o
to
E
<D
c
O)
a>
> zr.
Xi 00
CU
CU
o
00
■H
i-
CU
(U
<u
o
E
=3
(/)
>)
00
XJ •
Q.
•1- Q.
E CU
00
+->
•r—
c
i-
O
cu
a.
>
+->
<u
C
TC)
00
<u
O
OO
+->
=3 £=
fO
4->
cu
o
u
o
+->
Oi
o
x: •
C
CU
fO
•r—
fO
M-
o
(U
O)
fO
E
l-
4->
u
o. ^
+->
00 O
s-
t_
•a
+->
•r-
r—
-o
t/1
>>
cn
13
<_)
O
(U
00
C 1—
O) O
o
n3
c
i-
+->
n3
-o
x>
o
TCI
C
jcz
Q.
00 o
O
CU <D
sz <u
c
j:i
o
cu
O
(U
c
o
E
fO
•r-
Q_
to
fO +-»
s:
00 XJ
o
CJ
J-
c
CU
E
— J <U
O.
X
c
•O CU
r- E
<U -r-
•r- i_
U- O)
O.
X
c
o
•1—
T3 +-»
I— 03
CU >
•I- S-
U_ CU
00
^ i_ I—
CU to
r- >
cyt CU +->
,— Q_ CU ^
w ,— C\J
ed 03 00
C CU CJO
+J r- r-
rd 00 • — ' I — - —
J- CU •!-
CD oQ s:
o.
00
c
Q.
o
4/1
UJ
1—
I—
13
ra
>-
03
•1—
CJ
O-
O
CU
o
•r—
l-
o
LiJ
<
2:
r- E
^ — ' ro
=J CU
XI sz
M O
O
oa o
+-> CD
00 c cn
•r- CU I—
> I-
to rij
O Q.
>»
■♦->
•1—
cz
o
o
c
o
-M
CU
J-
cz
Z3
(0
-M
<_)
a.
23
o
l-
o
+-»
M
OO
191
I/)
or
>>
1
E
O)
o
r—
•1 —
Z3
00
4->
>>
00
r—
CZ
00
1— fO
00
<u
1
O)
o
«o
<u
i_
<U
>>+->-— 1
O)
T3
-o
•f—
a>
•
r—
s_
CD
<U
•1—
>>
c o
1 —
O)
+-»
•
•r-
<_>
i-
c:
in
O)
O
+-*
-M
Q.
a o c
S_
O
O)
=3 O
•r—
>
■»->
CL
03
<u
'f—
=3 a> T-
fO
o
fO
c:
<->
D.1—
lO
•p-
03
O
m
CL-a
00
Q.
o
fO
00
•r—
U3
+->
£=
s_
+-*
t/1
•r—
1/1
a; M-
0)
cz
03
_^
"O
o
O
CD
i_ -r-
i-
•a
O
Q.
00
U_
x:
Qi
o.
00
<U
x:
+->
(U
<D
M-
to
•H
u- -o •
03
o
c
O
00
Q.
00
c
00
c
•r-
V4-
o
03
o
(0
s-
E
•( —
3
XI
o
03
<4-
M-
to
1 —
^ O O
00
M-
c
1—
y-
>
i-
03
o
to
03
X5
UJ
J-
«3
(V (0 •
(U
o
o.
>>
(T3
•o
-a
<U
+-» in
+J
O)
-M
o
^
<D
>>
CD
-o
>>
C
c=
•1—
T3
i/i
a;
CD
>> <u 1
fT3
+J
■M
-M
>
o
C
X3
•1 —
•r—
03
o
e
<U
lo
^ 4-> O
X)
CZ
>)
C
o
(O
>
c
ro
Ol ra •
ro
jc
<D
<4-
-M
•TU
00
TO
00
»T3
O
o
4-
C
a>
o c:
a>
+->
r—
Q.
o
•r—
O
S_
J-
i-
>>
<X>
1 —
i_
o
i- .1-
CD
a.
CT
00
i_
Z3
a>
<1>
>>
cu
lO
T3
f—
00
<_>
o
c
J- E 3=
03
o
<U
<V
<u
C
Q.
o
c
+->
c
J3
q;
c
•r—
>> O Q.
o
i-
c
(U
<U
(U
03
<u
JC
<->
<T3
O
(— -o
•r—
Q
Q_ -O ^
M-
Q.
N
+->
o
00
o
CD
3
>
CD
C3l
(T3
O
1— CO
O) CO
sz <^
03 c—
3
c:
c
O
o
CD
CD
to
cn
+-»
r—
03
>
CO
00
>
*
<7i
00
(U
>
J3
x:
&.
O
03
ra
3Z
00
0)
03
03
(O
i-
E
i_
oo
'o)
o
rj
■H
E
+J
e'
00
>
+->
o
o
to
<u
(U
03
00
J-
o
s-
■o
00
-M
00
4-
o
Q
0)
cr
o
O
Q)
o
-M
c
c
o.
C
o
+->
00
c
00
+J
c
o
00
o.
o
•1—
(U
UJ
CO
i/i
00
S-
>»
s:
"
o
CD
CL.
■H
03
UJ
c
00
r~
•c—
■o
xz
00
ns
c
•r—
ra
l_
Q.
C
rcJ
a
i-
c
•t—
O
x:
o
O
=J
a>
O)
O)
0)
•r-
z:
o
Q
+->
E •
(/) x:
E
+->
t_
a>
c
>> E
UJ +->
•1—
i-
•r—
l-
T3
+->
E
o
x:
O
03
Q.
X
O)
C
o
•r—
X
Q.
(/) LI-
o
<u
Q-
ra
z
192
"O
+-»
CD
00
00
un
o
"O T-
c:
Q> ro
m
in
> i-
o.
o.
"O
• 1—
*~
QJ
X
4-* C_
*'
O •'
OJ
ID
QJ
(_) OJ
•1 — o
xi
Q)
CI
I/)
CD
o
'I"
<D
<L)
"O
TD
"O
+->
+->
+->
00
E •
O •
IT)
O >)
<D
O)
CD
<D
c=
fO
o in
i~ o
fO r—
-H
-M
XI
X2
o
OJ
Q. •
O
o
u
Qj ir>
>» fO
OJ
CD
4_>
J3 I
5-
1 — o
M- IT)
M-
<4-
CNJ
y-
"i
'e
U-
o
Ol
00 nD
o
fO c
n3
•(—
•1 —
UJ
cr m
O) <D
S- O)
■H
o
o c:
QJ
rj cn
>)0
>iO
T3
O
•r- zn.
c
-M o
CU
r3
Q.
<U CD
n3 Q. •
CD U->
<D un
Q. •
CX '
>
-a
•I- i-
+->
C O 00
00
00
O
o
i_
o
r- 3
c
i~ i~
i- 3
i- .
i_ .
OJ
s_
Z3 O
a> r—
CD
SZ SZ <1)
CD O
(D O
CD in
QJ in
00
Q.
Q. 1—
Q. -r-
OO
+->+->+->
> r—
> r—
XI
CD
O CL>
X ra
O C TCI
TZJ CD
"O CD
o o
o o
O
Q. X3
UJ u-
ro
<r XI
—1 +->
—I +->
CD
E
X5 -r-
C
"O CD
<- E
CD -r-
•r- i_
U- CD
a.
CNJ r-
r— r-
cn cr>
OO
E -r^
fO o
CD S-
00 h-
CD ^
x: CO
n3 CD
cd cri
c
00
o
iX)
Z3
r--
O
c
r~
03
o
C
+->
cn
OD
o
cn
00
cn
cn
00
1 —
+->
C30
03
>
13
CD
( —
51
S-
C
CD
sz
JZ
x:
«X>
oa
CO
1/1
00
>)
ro
to
OO
r-
>>
(XI
XI
CD
s:
QJ
CT>
o
*i
>
E
E
>
QJ
oe)
03
ra
x:
CD
CD
CD
03
03
OO
in
OQ
CQ
JC
q;
00
I-
d
OO
00
QJ
o
00
(0
■M
C
E
03
<D
CQ
00
i-
^ — >
r3
T3
4-*
U
13
00
XI
QJ
00
•(—
o
S-
(O
03
+-J
+J
00
00
O
jc
i_
•r-
4->
r3
C_)
^
13
CI
i-
•(—
Z3
13
rj
13
i_
■H
CD
rj
13
c
OO
E
o
h-
03
(O
O
i_
o
C
Cl_
rj
-M
CD
O
13
rj
o
OO
cn
00
s_
+->
i_
fO
o
OJ
i_
CD
to
+->
i-
:5
•1—
h-
f—
c
i-
E
o
E
1—
CD
>
00
CD
o
o
+->
O
O
QJ
o
J-
o
>
03
CD
O
J3
E
c:
E
c
E
^
>
o
CD
E
+->
-M
1
c
03
o
c
CI
03
o
-5^
03
03
fO
O
03
03
o
ro
o
fO
o
o
E
:E
Til
03
C/0
C
XI
C_J
o
fO
I/O
■»->
C/)
i_
CO
i_
I/O
X
i/^
_l
3
<:
00
OO
193
C •
*
on
•f— ^3
he
-M
+->
•( —
•1 —
O O
cn t/1
to
I/O
cr 03
J_
2
O
Qj
I/)
c
O.
O) o
(/I
+->
>> <U •
E
i_
O
o
■H
i- Q. +J
o
«^
* il
Q.
a;
•1 —
0)1/10
o
>>
I/O 0)
E
<4-
E
i
> c
a>
1— <u
<U -M
o
<-^-
•1 —
•1—
i_ •(-
■xj
■(-> cn
U1 o
o
UJ
r—
<L) CD +->
c c
03 03
O)
E ^ X
03 03
-a
TD
zr.
o +-> o;
ra
o i-
CD
Q.
a.
o o
•1 —
>i CD
>i
>
s_
M- zn
C (_)
i_
s_
i-
i_
xj $- E
Q)
•r- CL
03 03
03
O)
<U
<D
<u o
+->
C
E r-
E
>> +-> o
U
O) <u
Q.
o
O
<-t- a>
(13
•1-
C O)
J-
o
_j
_J
s: 03
OQ
»/1 +->
Q.
^ Z3
0 c
OD CT
01 03
<— s:
^oej CO
>> OD
O r— CTi
> OJ r—
sz
03 03
zn. a:
to
c= E
00
o
00
f— c
^ o
U3 00
r~ • ^ 3
cn I— o c
I — 03 CD
cn 03
-H f— s:
O oS CO
00 l_ >, CD
-O O) O) r— CT>
S- 1— > CU r—
fO r— i_ ^
> CU fO 03
00 i<i DC D::
03 •
+-> f—
to 03
<U
> 4->
•r- CU
cn
fO
03
4->
■M
OJ
CD
E
fO
V-
O
cu
s_
O
■o
4->
CD
CU
cn
00
cn
OJ
1 —
c:
x:
XI
03
fO
oo
zn.
r- 4->
C71 CU
' XJ
>» +-> vr>
CU 00 r~
s- CU cn
-a > f—
C T-
CU (U
o
XI CU
o >
i- 03
CU I—
C3L. 4+_
O U
r— i-
1— O)
(U O-
>- w
CO
00
cn
x:
u
i-
03
c
CU
o
00
00
CU
13
i-
CU
cr
•I—
CP
00
TU
ra
Z3
z:
O
4->
i_
to
CU
CU
E
+J
Z3
O
i-
C
C
=3
o
00
194
Goldstein et al. 1984b). The model was developed specifically to predict changes in
surface water hydrogen and aluminum ion concentrations because of their importance to
fish. It was divided into two modules: hydrological and chemical. The model incorpor-
ated the major flow pathways followed by precipitation and the major biogeochemical
processes that alter the chemical characteristics of water as it moves along these
pathways. The model accounted for the routing of the precipitation through the forest
canopy, soil horizons, and streams and lakes using mass-balance concepts and equations
which related flow to hydraulic gradients. The physical and chemical processes which
change the acid -base status of the water were simulated by rate and equilibrium expres-
sions. Mass balance transfers between gas, liquid, and solid phases were included in
the model. The aqueous constituents simulated were the major cations and anions,
aluminum and its complexes, organic acid analogues, and dissolved inorganic carbon.
Concentrations of free hydrogen ion or pH were derived from the solution alkalinity and
the total concentrations of inorganic carbon, organic acids, and monomeric aluminum.
The model was used to predict changes in acidity of Woods Lake (pH 4.7 to 5.1)
and Panther Lake (pH 5.3 to 7.8) given reductions in the total atmospheric sulphur
loads. A reduction in the incoming sulphur load by 50% was found to have little effect
on the pH of Panther Lake even after a 12 year simulation; in Woods Lake, however, the
pH increased substantially.
The model showed that movement of water through shallow or deep soil largely
determines the extent to which incident precipitation is neutralized as reported by Chen
et al. (1984) and Schofield (1984). Analysis of the two lake basins using the simulation
model and field data showed the watersheds to be net suppliers of base to the through-
flowing waters, and that the watershed internally provided two-thirds of the amount of
the atmospheric acidity naturally.
The ILWAS model can be criticized for treating the role of internally generated
organic acids simpl i sti cal 1 y . However, it is important to note that Schofield (1984)
observed that organic acids were not important in determining the pH of the lakes studied
in the ILWAS program.
It is also important to note that while the two study lakes were morphometri-
cally similar, the model correctly identified the effects of the differences in soil
characteristics and the resultant effect that would have on long term pH changes. Woods
Lake is surrounded by thin till which is less permeable than that of the deeper tills
surrounding Panther Lake. Thus, a large fraction of the throughflow water draining into
Woods Lake moves through organically rich, shallow soil horizons, whereas simi lar waters
in the Panther Lake drainage basin are in contact with base rich tills for a longer
period. Therefore, the lateral flow waters entering Woods Lake after moving through the
shallow organic horizons result in low pH and low alkalinity values. Removal of the
incoming acid loading to the lake, which Gherini et al . (1984) had estimated to account
for up to 60% of the total basin acidity would therefore allow the pH in Woods Lake to
rise; conversely, the pH of the already well buffered waters of the Panther Lake Basin
would not.
195
8.10 EFFECTS OF ACIDIC DEPOSITION ON SURFACE WATER ACIDIFICATION: LITERATURE CITED
Alfheim, I., M. Stobet, N. Gjos, A. Bjorseth, and S. Wilhelmsen. 1980. An analysis of
organic micropol 1 utants in aerosols. Ijn: Ecological Impact of Acid
Precipitation, Proceedings of an International Conference, eds. D. Drablos and
A. Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project. Oslo, Norway;
pp. 100-101.
Allen H.L. 1972. Phytoplankton photosynthesis, micronutrient interactions, and inorganic
carbon availability in a soft-water Vermont lake. In: Nutrients and
Eutrophication , the Limiting-Nutrient Controversy, ed . G.E. Likens. Special
Symposium, Vol. 1. American Society of Limnology and Oceanography. Lawrence,
Kansas, pp. 68-83. (Original not seen; information taken from Marcus et. al.
1983.)
Aimer, B., W. Diskson, C. Ekstrom, and E. Hornstrom. 1978. Sulfur pollution and the
aquatic ecosystem. Vn: Sulfur in the Environment. Part II, ed . J. Nriagu.
New York: John Wiley and Sons, pp. 273-311.
Aimer, B., W. Dickson, C. Ekstorm, E. Hornstorm, and U. Miller. 1974. Effects of acidi-
fication on Swedish lakes. Ambio 3: 30-36.
Baker, J. P. 1983a. 5.5 Plankton. In: Chapter E-5. Effects on Aquatic Biology. The
Acidic Deposition Phenomenon and Its Effects. Critical Assessment Review
Papers. Volume II - Effects Sciences. Public Review Draft, eds. A. P. Altshuller
and R.A. Linthurst. US Environmental Protection Agency Report 600/8-83-01 6B .
pp. 5-45 to 5-76.
Baker, J. P. 1983b. 5.6 Fishes. In: Chapter E-5. Effects on Aquatic Biology. The Acidic
Deposition Phenomenon and Its Effects. Critical Assessment Review Papers.
Volume II - Effects Sciences. Public Review Draft, eds. A. P. Altshuller and
R.A. Linthurst. US Environmental Protection Agency Report 600/8-83-01 6B.
pp. 5-76 to 5-137.
Beamish, R.J. 1976. Acidification of lakes in Canada by acid precipitation and the
resulting effects on fish. Water, Air, and Soil Pollution 6: 501-514.
Beamish, R.J. 1974. Loss of fish populations from unexploited remote lakes in Ontario,
Canada as a consequence of atmospheric fallout of acid. Water Research 8:
85-95.
Beamish, R.J. 1972. Lethal pH for the white sucker Catostomus commersoni (Lacepede).
Transactions of the American Fisheries Society 101: 355-358.
Beamish, R.J. and H.H. Harvey. 1972. Acidification of the LaCloche Mountain lakes,
Ontario and resulting fish mortalities. Journal of the Fisheries Research
Board of Canada 29: 1131-1143.
Beamish, R.J., W.L. Lockhart, J.C. Van Loon, and H.H. Harvey. 1975. Long-term acidifica-
tion of a lake and resulting effects on fishes. Ambio 4: 98-102.
Bell, 6.R. 1971. Effect of low pH on the survival and emergence of aquatic insects. Water
Resources 5: 313-319.
Bell, H.L. and A.V. Nebeker. 1969. Primary studies on the tolerance of aquatic insects
to low pH. Journal of the Kansas Entomological Society 42(2): 230-237.
Best, M.D. and J.H. Peverly. 1981. Water sediment chemistry and elemental composition of
macrophytes in thirteen Adirondack lakes (abstract). In: Proceedings of
International Symposium on Acidic Precipitation and Fishery Impacts in North
Eastern North America, North Eastern Division American Fisheries Society. 1981
August 2-5; Cornell University, Ithaca, New York; p. 348.
Bjarnborg, B. 1983. Dilution and acidification effects during the spring flood of four
Swedish mountain brooks. Hydrobiologia 101: 19-26.
196
Borgstrom, R. and G.R. Hendrey. 1976. pH tolerance of the first larval stages of Lepi-
durus arcti cus (Pallas) and adult Gammarus lacustri s G.O. Sars. SNSF Project,
Internal Report 22/76. Oslo-As, Norway. 37 pp.
Braekke, F.H. 1981. Hydrochemi stry in low-pH-soils of South Norway. 1. Peat and soil
water quality. Meddelelser fra Norsk Institutt for Skogf orskni ng 36(11): 1-32.
Brezonik, P.L., E.S. Edgerton, and CD. Hendry. 1980. Acid precipitation and sulfate
deposition in Florida. Science 208: 1027-1029.
Brock, T.D. 1978. Thermophilic Organisms and Life at High Temperatures. New York:
Springer-Verlag. 465 pp.
Cadle, S.H., J.M. Dasch, and N.E. Grossnickle. 1984. Retention and release of chemical
species by a northern Michigan snowpack. Water, Air, and Soil Pollution 22:
303-319.
Canfield, D.E., Jr. 1983. Sensitivity of Florida lakes to acidic precipitation. Water
Resources Research 19(3): 833-839.
Chen, C.W., S.A. Gherini, N.E. Peters, P.S. Murdoch, R.M. Newton, and R.A. Goldstein.
1984. Hydrologic analyses of acidic and alkaline lakes. Water Resources
Research 20(12): 1875-1882.
Costa, H.H. 1967. Responses of Gammarus pulex (L.) to modified environment II. Reactions
to abnormal hydrogen ion concentrations. Crustacaena 13: 1-10.
Davis, P. and B. Ozburn. 1969. The pH tolerance of Daphnia pulex (Lydig, emend.,
Richard.) Canadian Journal of Zoology 47: 1173-1175.
Davis, R., M. Smith, J. Bailer, and S. Norton. 1978. Acidification of Maine (USA) lakes
by acidic precipitation. Verhandl ungen Internationale Vereinigung fiir
Limnologie 20: 532-537.
Dickson, W. 1980. Properties of acidified waters. In: Ecological Impact of Acid Pre-
cipitation, Proceedings of International Conference, eds. D. Drablos and A.
Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway;
pp. 75-83.
Dillon, P.J., D.S. Jeffries, and W.A. Scheider. 1982. The use of calibrated lakes and
watersheds for estimating atmospheric deposition near a large point source.
Water, Air, and Soil Pollution 18: 241-258.
Dillon, P.J., D.S. Jeffries, W. Snyder, R. Reid, N.D. Yan, D. Evans, J. Moss, and W.A.
Scheider. 1978. Acidic precipitation in south-central Ontario: recent obser-
vations. Journal of the Fisheries Research Board of Canada 35: 809-815.
Drablos, D. and A. Tollan, eds. 1980. Ecological impact of acid precipitation. Proceed-
ings of an International Conference. 1980 March 11-14; Sandefjord, Norway.
SNSF Project, Oslo, Norway; 383 pp.
Driscoll, C.T. 1980. Aqueous speciation of aluminum in Adirondack region of New York
State, U.S.A. In: Ecological Impact of Acid Precipitation, Proceedings of an
International Conference, eds. D. Drablos and A. Tollan; 1980 March 11-14,
Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 214-215.
Driscoll, C.T. and R.M. Newton. 1985. Chemical characteristics of Adirondack lakes.
Environmental Science and Technology 19(11): 1018-1024.
Edmondson, W.T. 1944. Ecological studies of sessile rotatoria. Ecological Monographs
14: 31-66.
Eilers, J.M., G.E. Glass, K.E. Webster, and J. A. Rogalla. 1983. Hydrologic control of
lake susceptibility to acidification. Canadian Journal of Fisheries and Aquatic
Sciences 40: 1896-1904.
197
Elzerman, A.W. 1983. Effects of acid deposition (rain) on a piedmont aquatic ecosystem:
acid inputs, neutralization, and pH changes. Final Technical Completion Report
B-141 -SC to Bureau of Reclamation, US Department of the Interior, Washington,
DC. 79 pp.
Flora, M.D. and P.C. Rosendahl. 1982. The impact of atmospheric deposition on the water
quality of Everglades National Park. Proceedings of the American Water Resources
Association International Symposium on Hydrometeorology , eds. R. Herrmann and
A.I. Johnson. 1982 June; pp. 55-61.
Gahnstrom, G., G. Andersson, and S. Fleischer. 1980. Decomposition and exchange processes
in acidified lake sediment. In: Ecological Impact of Acid Precipitation,
Proceedings of an International Conference, eds. D. Drablos and A. Tollan.
1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 306-307.
Galloway, J.N., G.E. Likens, and M.E. Hawley. 1984. Acid precipitation: natural vs.
anthropogenic components. Science 226: 829-831.
Galloway, J.N., C.L. Schofield, G.R. Hendrey, N.E. Peters, and A.H. Johannes. 1980.
Sources of acidity in three lakes acidified during snowmelt. Xn^: Ecological
Impact of Acid Precipitation, Proceedings of International Conference, eds.
D. Drablos and A. Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project,
Oslo, Norway; pp. 264-265.
Galloway, J.N., E. Cowling, E. Gorham, and W. McFee. 1978. A national program for
assessing the problem of atmospheric deposition (acid rain). Report, Council
of Environmental Quality, Fort Collins, Colorado.
Gherini, S.A., C.W. Chen, L. Mok, R.A. Goldstein, R.J.M. Hudson, and G.F. Davis. 1984.
The ILWAS Model: Formulation and Application in the Integrated Lake-Watershed
Acidification Study, Volume 4. Summary of Major Results. Electric Power
Research Institute. EA-3221 . pp. 7-1 to 7-45.
Gibson, J.H., J.N. Galloway, C. Schofield, W. McFee, R. Johnson, S. McCarley, N. Dise,
and D. Herzog. 1983. Rocky Mountain acidification study. US Fish and Wildlife
Service, Division of Biological Services, Eastern Energy and Land Use Team.
Report FWS/OBS-80/40.17. 137 pp.
Glass, G.E. 1984. Susceptibility of aquatic and terrestrial resources of Minnesota,
Wisconsin, and Michigan to impacts from acid precipitation: informational
requirements. US Environmental Protection Agency Report EPA 600/D-84-064 .
Available from National Technical Information Services as NTIS PB84-174929.
pp. 21-22.
Glass, G.E., T.G. Brydges, and O.E. Loucks, eds. 1981. Impact assessment of airborne
acidic deposition on the aquatic environment of the United States and Canada.
US-Canada Memorandum of Intent on Transboundary Air Pollution. Interim Report.
206 pp.
Goldstein, R.A., C.W. Chen, and S.A. Gherini. 1984a. Integrated Lake-Watershed Acidifi-
cation Study - Summary. In: The Integrated Lake-Watershed Acidification
Study. Volume 4. Summary of Major Results, eds. R.A. Goldstein and S.A.
Gherini. Electrical Power Research Institute EA-3221. 1-1 to 1-15 pp.
Goldstein, R.A., S.A. Gherini, C.W. Chen, L. Mok, and R.J. Hudson. 1984b. Integrated
acidification study (ILWAS): a mechanistic ecosystem analysis. Philosophical
Transactions of the Royal Society of London B 305: 409-425.
Gorham, E. and W.W. McFee. 1980. Effects of acid deposition upon outputs from terrestrial
to aquatic ecosystems. In: Effects of Acid Precipitation on Terrestrial
Ecosystems, eds. T.C. Hutchinson and M. Havas. NATO Conference Series 1:
Ecology, Volume 4. New York: Plenum Press, pp. 465-480.
Grahn, 0. 1977. Macrophyte succession in Swedish lakes caused by deposition of airborne
acid substances. Water, Air, and Soil Pollution 7: 295-305.
Griffiths, D. 1973. The structure of an acid moorland pond community. Journal of Animal
Ecology 42: 263-283.
198
Haines, T.A. 1981. Acidic precipitation and its consequences for aquatic ecosystems: a
review. Transactions of the American Fisheries Society 110: 669-707.
Haines, T.A., J. Akielaszek, and P. Rago. 1983. A regional survey of the chemistry of
headwater lakes and streams in New England: vulnerability to acidification.
Fish and Wildlife Service, US Department of the Interior. Air Pollution and
Acid Rain Report No. 15. FWS/OBS-80/40 . 1 5 . Available from National Technical
Information Service as PB84-103357. 141 pp.
Hall, R.J., G.G. Likens, S.B. Fiance, and G.R. Hendrey. 1980. Experimental acidification
of a stream in the Hubbard Brook Experimental Forest, New Hampshire. Ecology
61(4): 976-989.
Hargreaves, J.W., E.J.H. Lloyd, and B.A. Whitton. 1975. Chemistry and vegetation of
highly acidic streams. Freshwater Biology 5: 563-576.
Harte, J., G.P. Lockett, R.A. Schneider, H. Michaels, and C. Blanchard. 1985. Acid pre-
cipitation and surface-water vulnerability on the western slope of the high
Colorado Rockies. Water, Air, and Soil Pollution 25: 313-320.
Harvey, H.H. 1980. Widespread and diverse changes in the biota of North American lakes
and rivers coincident with acidification. Iri: Ecological Impact of Acid
Precipitation, Proceedings of an International Conference, eds. D. Drablos and
A. Tollan. 1980 Marcy 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway;
pp. 93-98.
Hendrey, G. 1976. Effects of low pH on the growth of periphytic algae in artificial
stream channels. SNSF Project Report IR 2576, Oslo, Norway.
Hendrey, G.R., J.N. Galloway, S.A. Norton, C.L. Schofield, D.A. Burns, and P.W. Schaffer.
1980a. Sensitivity of the Eastern United States to acid precipitation impacts
on surface waters. In: Ecological Impact of Acid Precipitation, Proceedings
of an International Conference, eds. D. Drablos and A. Tollan. 1980 March
11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 216-217.
Hendrey, G.R., J.N. Galloway, and C.L. Schofield. 1980b. Temporal and spatial trends in
the chemistry of acidified lakes under ice cover. In: Ecological Impact of
Acid Precipitation, Proceedings of an International Conference, eds. D. Drablos
and A. Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo,
Norway; pp. 266-267.
Herczeg, A.L., W.S. Broecker, R.F. Anderson, S.L. Schiff, and D.W. Schindler. 1985. A
new method for monitoring temporal trends in the acidity of fresh waters.
Nature 315: 133-135.
Hornbeck, J.W., G.E. Likens, and J.S. Eaton. 1977. Seasonal patterns in acidity of
precipitation and their implications for forest stream ecosystems. Water, Air,
and Soil Pollution 7: 355-365.
Hultberg, H. 1977. Thermally stratified acid water in late winter - a key factor inducing
self-accelerating processes which increase acidification. Water, Air, and Soil
Pollution 7: 279-294.
Hunt, D. and C.E. Boyd. 1981. Alkalinity losses from ammonia fertilizers used in fish
ponds. Transactions of the American Fisheries Society 110: 81-85.
Hutchinson, G.E., W. Gizyn, M. Havas, and V. Zobens. 1978. Effect of long-term ligmite
burns on arctic ecosystems at the Smoking Hills, NWT. In: Trace Substances in
Environmental Health - XII, ed. D.D. Hemphill. Columbia, Missouri: University
of Mi ssouri .
Jeffries, D.S., CM. Cox, and P.J. Dillon. 1979. The depression of pH in lakes and
streams in central Ontario during snowmelt. Journal of the Fisheries Research
Board of Canada 36: 640-646. (Original not seen; information taken from Marcus
et al. 1983.)
Jensen, K. and E. Snekvik. 1972. Low pH levels wipe out salmon and trout populations in
southernmost Norway. Ambio 1: 223-225.
199
Johannes, A.H. and E.R. Altwicker. 1980. Atmospheric inputs to three Adirondack lake
watersheds. In: Ecological Impact of Acid Precipitation, Proceedings of an
International Conference, eds. D. Drablos and A. Tollan. 1980 March 11-14;
Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 256-257.
Johannessen, M., A. Skartveit, and R.F. Wright. 1980. Streamwater chemistry before,
during, and after snowmelt. In: Ecological Impact of Acid Precipitation,
Proceedings of an International Conference, eds. D. Drablos and A. Tollan.
1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 224-225.
Johnson, A.H. 1979a. Acidification of headwater streams in the New Jersey pine barrens.
Journal of Environmental Quality 8(3): 383-386.
Johnson, A.H. 1979b. Evidence of acidification of headwater streams in the New Jersey
pinelands. Science 206: 834-836.
Jones, H.G., M. Ouellet, and D.F. Braekke. 1980. The evolution of acidity in surface
waters of Laurentides Park (Quebec, Canada) over a period of 40 years. In:
Ecological Impact of Acid Precipitation, Proceedings of an International
Conference, eds. D. Drablos and A. Tollan. 1980 March 11-14; Sandefjord,
Norway; SNSF Project, Oslo, Norway; pp. 226-227.
Kaplan, E., H.C. Thude, and A. Protas-Rocks . 1981. Soils and water quality: relationships
and implications for effects of acid precipitation on surface water in the
northeastern United States. Environmental Science and Technology 15(5):
539-544.
Keller, W., J. Gunn, and N. Conroy. 1980. Acidification impacts on lakes in the Sudbury,
Ontario, Canada area. In: Ecological Impact of Acid Precipitation, Proceedings
of an International Conference, eds. D. Drablos and A. Tollan. 1980 March
11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 228-229.
Kling, G.W. and M.C. Grant. 1984. Acidic precipitation in the Colorado front range: an
overview with the prediction for significant events. Arctic and Alpine
Research 16(3): 321-329.
Kramer, J. 1976. Geochemical and lithological factors in acid precipitation. USDA Forest
Service General Technical Report NE-23. pp. 611-618.
Krug, E.C. and C.R. Frink. 1983a. Acid rain on acid soil: a new perspective. Science
221: 520-525.
Krug, E.C. and C.R. Frink. 1983b. Effects of acid rain on soil and water. Bulletin 811.
The Connecticut Agricultural Experiment Station, New Haven. November 1983. 45
pp.
Laake, M. 1976. Effekter au lav pH pa produksjon, nedbrytning og stof f kretslop i lit-
toralsonen. SNSF Project. Internal Report 29/76. Oslo-As, Norway.
LaChance, M., B. Bobee, and Y. Grimard. 1985. Sensitivity of southern Quebec lakes to
acidic precipitation. Water, Air, and Soil Pollution 25: 115-132.
Leivestad, H., G. Hendrey, I. P. Muniz, and E. Snekvik. 1976. Effects of acidic precipi-
tation on freshwater organisms. Ijn: Impact of Acid Precipitation on Forest
and Freshwater Ecosystems in Norway. SNSF Research Report No. 6, ed. F.H.
Braekke. Oslo, Norway: SNSF Project, pp. 87-111.
Lewis, W. 1982. Changes in pH and buffering capacity of lakes in the Colorado Rockies.
Limnology and Oceanography 27: 167-172.
Liaw, W.K. 1982. Chemical sensitivity of Northern Saskatchewan waters to acidification.
in: Proceedings of Symposium/Workshop on Acid Forming Emissions in Alberta and
Their Ecological Effects. Alberta Department of the Environment, Canadian
Petroleum Association, Oil Sands Environmental Study Group; 1982 March 9-12;
Edmonton, Alberta; pp. 485-509.
Likens, G.E., F.H. Bormann, R.S. Pierce, J.S. Eaton, and N.M. Johnson. 1977. Biogeo-
chemistry of a Forested Ecosystem. New York: Spri nger-Verlag . 146 pp.
200
Lunde, G.J., N.G. Gether, and M. Lande. 1977. Organic mi c ropoll utants in precipitation
in Norway. Atmospheric Environment 11: 1007-1014.
Magnuson, J.J. 1983. 5.1 Introduction. I_n: Chapter E-5. Effects on Aquatic Biology.
Ihe Acidic Deposition Phenomenon and Its Effects. Critical Assessment Review
Papers. Volume II - Effects Sciences. Public Review Draft, eds. A. P. Altshuller
and R.A. Linthurst. US Environmental Protection Agency Report 600/8-83-01 6B.
pp. 5-1 to 5-3.
Malley, D.F. 1980. Decreased survival and calcium uptake by the crayfish Oronectes
vi ri 1 i s in low pH. Canadian Journal of Fisheries and Aquatic Sciences 37:
364-37 2.
Marcus, M.D., B.R. Parkhurst, and F.E. Payne. 1983. An assessment of the relationship
among acidifying depositions, surface water acidification, and fish populations
in North America. Volume 1. Prepared by Western Aquatics, Inc. for Electric
Power Research Institute, as EA-3127, Research Project 1910-2. 105 pp.
Mashiko, K., K. Jozuka, and K. Asakura. 1973. Different types of chloride cells in the
gills of Tri bol odon hakonensi s from Lake Osoresan-ko. Ann. Rep. Noto. Marine
Lab 13: 33-37. XH- Exceptional Tolerance to Low pH of Some Tropical Blackwater
Fish, eds. W.A. Dunson, F. Swarts, and M. Silvestri. Journal of Experimental
Zoology 201 : 1 57 -1 62.
Mason, J. and H.M. Seip. 1985. The current state of knowledge on acidification of surface
waters and guidelines for further research. Ambio 14: 45-51.
Melack, J.M., J.L. Stoddard, and D.R. Dawson. 1982. Acid precipitation and buffer
capacity of lakes in the Sierra Nevada, California. Proceedings of the American
Water Resources Association International Symposium on Hydrometeorol ogy , eds.
R. Herrmann and A.I. Johnson; 1982 June; pp. 35-41.
Miller, G.E., I. Wile, and G.G. Hitchin. 1982. Patterns of accumulation of selected
metals in members of the soft -water macrophyte flora of central Ontario lakes.
Aquatic Botany 15: 53-64.
Mills, K.H. 1982. Fish population responses during the experimental acidification of a
small lake. Presented at American Chemical Society Meeting, Las Vegas, Nevada,
1982 March.
Morgan, D. 1984. Acidification of headwater streams in the New Jersey pinelands. A
re-evaluation. Limnology and Oceanography 29(6): 1259-1266.
Mount, D.I. 1973. Chronic effect of low pH on fathead minnow survival, growth and repro-
duction. Water Resources 7: 987-993.
National Wildlife Federation. 1984. Acid rain: Its state by state impacts. Washington,
DC. Available from National Technical Information Service as NTIS PB84-198332.
101 pp.
Norton, S.A. 1980. Geological factors controlling the sensitivity of aquatic ecosystems
to acidic precipitation. I_n: Atmospheric Sulfur Deposition: Environmental
Impact and Health Effects, eds. O.S. Shriner, C.R. Richmond, and S.E. Lindberg.
Ann Arbor, Michigan: Ann Arbor Science, pp. 521-529. (Original not seen;
information taken from Marcus et al . 1983.)
Norton, S., R. Davis, D. Brakke, D. Hanson, K. Kenlan, and P. Sweets. 1981. Responses
of northern New England Lakes to atmospheric inputs of acid and heavy metals.
Completion Report, Project A-048-ME, Office of Water Research and Technology,
Washington, DC. 90 pp.
Oakland, K.A. 1980. Okologi og utbredelse til Gammarus lacustric G. 0. Sars i Norge, med
vekt pa f orsuringsproblemer. SNSF Project, Internal Report 67/80. Oslo-As,
Norway.
Overrein, L.N., H.M. Seip, and A. Tollan. 1980. Acid precipitation-effects on forests
and fish. Final Report of the SNSF-Project 1972-1980. Research Report 19.
Aas, Norway. 175 pp.
201
Panel on Processes of Lake Acidification: Environmental Studies Board; Commission on
Physical Sciences, Mathematics, and Resources. 1984. Acid deposition: pro-
cesses of lake acidification. Available from the National Technical Information
Service as PB84-216175. 11 pp.
Parent, S. and R. Cheetham. 1980. Effects of acid precipitation on Daphnia magna .
Bulletin of Environmental Contamination and Toxicology 25: 298-304.
Pennak, R.W. 1978. Fresh-water Invertebrates of the United States. New York: John Wiley
and Sons. 803 pp.
Pfeiffer, M. and P. Festa. 1980. Acidity status of lakes in the Adirondack Region of New
York in relation to fish resources. New York Department of Environmental
Conservation Report FW-P168. Albany, NY. 36 pp.
Rahel, F.J. and J.J. Magnuson. 1983. Low pH and the absence of fish species in naturally
acidic Wisconsin lakes: inferences for cultural acidification. Canadian
Journal of Fisheries and Aquatic Sciences 40: 3-9.
Rosenqvist, I.Th. 1978a. Acid precipitation and other possible sources for acidification
of rivers and lakes. The Science of the Total Environment 10: 271-272.
Rosenqvist, I.Th. 1978b. Alternative sources for acidification of river water in Norway.
The Science of the Total Environment 10: 39-49.
Scheider, W.A., J. Jones, and B. Cave. 1976. A preliminary report on the neutralization
of Nelson lake near Sudbury, Ontario. Ontario Ministry of the Environment.
Rexdale, Ontario.
Schindler, D.W. 1980. Experimental acidification of a whole lake: A test of the oligo-
trophication hypothesis. Ijn: Ecological Impact of Acid Precipitation,
Proceedings of an International Conference, eds. D. Drablow and A. Tollan.
1980 March 11-14; Sandefjord, Norway; SNSF Project, Oslo, Norway; pp. 370-374.
Schofield, C.L. 1984. Surface water chemistry in the ILWAS basins. In: The Integrated
Lake-Watershed Acidification Study. Volume 4. Summary of Major Results.
Electric Power Research Institute, Report EPRI EA-3221 , pp. 6-1 to 6-31.
Schofield, C.L. 1982. Historical fisheries changes as related to surface water pH changes
in the United States. In: Acid Rain/Fisheries. Proceedings of an International
Symposium on Acidic Precipitation and Fishery Impacts in Northeast North
America, ed . R.E. Johnson. Ithaca, NY; pp. 57-67. (Original not seen;
information taken from Havas et al . 1982.)
Schofield, C.L. 1976a. Acid precipitation: effects on fish. Ambio 5(5-6): 228-230.
Schofield, C.L. 1976b. Lake acidification in the Adirondack mountains of New York:
causes and consequences. Ijn: Proceedings of the First International Symposium
on Acid Precipitation and the Forest Ecosystem. US Department of Agriculture,
USDA Forest Service General Technical Report NE-23, eds. L.S. Dochinger and
T.A. Seliga. 1975 May 12-15; Columbus, Ohio; p. 477.
Scruton, D.A. 1983. A survey of headwater lakes in insular Newfoundland, with special
reference to acid precipitation. Canadian Technical Report of Fisheries and
Aquatic Sciences, No. 1195. 110 pp.
Seip, H.M. 1980. Acid snow-snowpack chemistry and snowmelt. Ijn: Effects of Acid Pre-
cipitation on Terrestrial Ecosystems, eds. T.C. Hutchinson and M. Havas. NATO
Conference Series 1: Ecology, Volume 4. New York: Plenum Press, pp. 77-94.
Shaffer, P.W. and J.N. Galloway. 1982. Acid precipitation: the impact on two headwater
streams in Shenandoah National Park, Virginia. Proceedings of the American
Water Works Association International Symposium on Hydrometeorol ogy , eds. R.
Herrmann and A.I. Johnson. 1982 June; pp. 43-53.
Smith, R.F. 1957. Lakes and ponds. Fishery survey report number 3. New Jersey Depart-
ment of Conservation and Economic Development, Division of Fish and Game.
Trenton, New Jersey. 198 pp.
202
Somers, K.M. and H.H. Harvey. 1984. Alteration of fish communities in lakes stressed by
acid deposition and heavy metals near Wawa, Ontario. Canadian Journal of
Fisheries and Aquatic Sciences 41: 20-29.
Strachan, W. and H. Huneault. 1979. Polychlorinated biphenyls and organochl ori ne pesti-
cides in Great Lakes precipitation. Journal of Great Lakes Research 5: 61-68.
Sutcliffe, D.W. and T.R. Carrick. 1973. Studies on mountain streams in the English Lake
District. Freshwater Biology 3: 437-462.
Svardson, G. 1976. Interspecific poulation dominance in fish communities of Scandinavian
lakes. Report of the Institute of Freshwater Research at Drottni ngholm 55:
144-171 .
Telang, S.A. 1987. Surface Water Acidification Literature Review. Prep for the Acid
Deposition Research Program by the Kananaskis Centre for Environmental
Research. The University of Calgary, Calgary, Alberta. ADRP-B-01 -87 . 123 pp.
Tetra Tech Inc. 1984. The Integrated Lake-Watershed Acidification Study. Volume 3. Lake
chemistry program. Electric Power Research Institute Report, EPRI EA-3221 , pp.
7-1 to 7-7.
Tetra Tech Inc. 1983 and 1984. Volumes 1 to 4. The Integrated Lake -Watershed Acidifi-
cation Study. Electric Power Research Institute Report, EPRI EA-3221.
Tomlinson, G., R. Brouzes, R. McLean, and J. Kadlecek. 1980. The role of clouds in
atmospheric transport of mercury and other pollutants. I_n: Ecological Impact
of Acid Precipitation, Proceedings of an International Conference, eds. D.
Drablos and A. Tollan. 1980 March 11-14; Sandefjord, Norway; SNSF Project,
Oslo, Norway; pp. 134-137.
Tonnessen, K.A. and J. Harte. 1982. Potential for acid precipitation damage to lakes of
the Sierra Nevada, California. I_n: Proceedings of the American Water Resources
Association. International Symposium on Hydrometeorology , eds. R. Herrmann and
A.I. Johnson. 1982 June; pp. 75-78.
Trojnar, J.R. 1977. Egg and larval survival of white suckers Catostomus comnersoni at
low pH. Journal of the Fisheries Research Board of Canada 34: 574-579.
Turk, J.T. and D.B. Adams. 1983. Sensitivity to acidification of lakes in the Flat Tops
Wilderness Area, Colorado. Water Resources Research 19: 346-350.
Ueno, M. 1958. The disharmonious lakes of Japan. Verhand 1 ungen Internationale
Vereinigung-Limnologie 13: 217-226.
Watt, W.D., CD. Scott, and S. Ray. 1979. Acidification and other chemical changes in
Halifax county lakes after 21 years. Limnology and Oceanography 24(6):
1154-1161 .
Wright, R.F. 1983. Input -output budgets at Langtjern, a small acidified lake in southern
Norway. Hydrobiologia 101: 1-12.
Yan, N. and P. Stokes. 1978. Phytoplankton of an acidic lake, and its responses to
experimental alterations of pH. Environmental Conservation 5: 93-100.
PART II.
ACIDIC DEPOSITION IN THE ALBERTA CONTEXT
203
9. ACIDIC DEPOSITION IN THE ALBERTA CONTEXT
9.1 MAJOR BIOPHYSICAL COMPONENTS OF ALBERTA
The major biophysical components of Alberta have been reviewed by Jaques (1987).
These include 1:1,000,000 mapping of macrocl imatic regions, physiographic regions,
physiographic subdivision, bedrock geology, and vegetation for the province with the
exception of the National Parks areas. Each bedrock formation in Alberta has been rated
according to its potential buffering capacity with respect to acid deposition at a scale
of 1:1,000,000. Rocks with low to no buffering capacity occur where relatively pure
siliceous rocks of sedimentary, igneous, and metamorphic origin are found, and these are
mainly in the northeastern portion of the province and along the Rocky Mountains and
their foothills. The map also indicates the distribution of unlimited, high to medium,
and medium to low buffering capacity in bedrock formations.
The twelve macrocl imatic regions of the province recognized by Strong and
Leggat (1981) are also mapped and the dominant vegetative species complexes within each
have been identified. Jaques (1987) recognized 134 vegetation community types throughout
Alberta. The community type analysis provided by Jaques (1987) shows that major
cl imatological ly controlled subdivisions occur in the province. These are: Subalpine,
Montane, Parkland, and Boreal Uplands areas. Jaques suggests that similar subdivisions
occur in other regions of the province but to date these have not been documented.
The sensitivity of macrod imatic and/or vegetation community types to acidic
deposition was not provided in the documentation because of a lack of sufficient data on
which to base such predictions at this time (Jaques 1987).
9.2 ACIDIC DEPOSITION AND ALBERTA FORESTS
Research into the effects of acidic deposition on forest ecosystems in Alberta
is extensive and has been documented briefly in Table 38. These programs indicate that
as a result of acidic deposition, soil acidification with accompanying solubilization of
aluminum and manganese is becoming evident in some areas (Baker et al. 1977; Nyborg
et al. 1977; Addison and Puckett 1980; and Addison 1984). Lore (1984), however, found
no evidence of change in soil pH downwind from a sour gas plant near Pincher Creek,
Alberta. The soils on which this study was conducted were of the loam to clay-loam
variety, not likely to be sensitive to acidification. Studies such as those of Legge
et al. (1977), Legge (1982), and Addison et al. (1984) have shown convincingly that
photosynthetic capacity of major tree species can be reduced close to emission sources.
These studies also indicate that there have been measurable effects at the biochemical
level such as chlorophyll destruction and altered energy allocation in trees in close
proximity to emission sources (Malhotra 1977; Harvey and Legge 1979). The literature
also indicates that, overall, forest growth may be reduced as a result of acidic
deposition in Alberta near point sources (Legge et al. 1984, 1986; Amundson et al.
1986). Non-point source contributions as a result of farming practices also contribute
to acidification. Sanderson (1984) pointed out that agricultural fertilizers contribute
approximately 25 times more to soil acidity on a yearly basis than does atmospheric
deposition. This implies that fertilizers contribute more to potential soil acidity in
Alberta than do industrial sources. An indication of the seriousness of this problem
204
<T3 ra
-o -o
C CL>
<4- C
E -r- •
+-> r— O)
in <t -t-
<-l-
-O (D
C r— -O
=) +->
o -o -1-
•r- O O
l_) (/I 00
O)
i- 0)
Q. • ^
C -M
-o O
<D ..- S_
>— 4-> O)
O. =3 S_
e r- to
(O O O)
c/1 -^r
•I- 03
TD CD
rO —I
r— CD
Q. ^ C
o -M s: >
S- nD O
<4- > -O O
a; c
4-> r- to
■o -o Q. x:
fO
4-> o 3 -o
c o o o
CD CM — I +->
E V ro
O 00 O OJ
CM
^-> to
O • i- O)
O) OJ <U t/l
14-1 — CL m
y- -r- Q. O
LlJ Q. 3 S
•1- C
l/l O
O -r-
Q. +J
<D -i-
■O I/O
> s_ re)
O <D S-
(_) > CD
O
<D (_) (N
O O
3 -O (/I
Q. U Q.
t/> 3 CD
-O O)
E
CD
> O
CD O
O r- -O
Q. l/l (O
fO l/l f —
-o 3
<N C l/l O
O rO O) CD l/l
l/l c
l/l C •!— O)
^ CD O -— >
E -r- Q. -r- .r-
Q. <J l/l fO -H
D. <D CD +-> (U
Q. q; C
l/l
CO
O CD
l/l l/l
>> s-
03 CD
■O 14-
+-> l/l C i-
■H -O CD •!- -O
(tJ CD I— C
CD <_) Xi "O O
•1- :3 •!- •!- CL
E "d l/l Q. l/l
3 CD •!- fO CD
3 - -
O <- •■ CD O) <a
I — x: o (_) <D
1— CD (_) <- 3 3 -(->
•r- -O S- <D i- i-
S -- •>- C
Q. Q. S-
CD ra
CD CD CD ^ -H O S-
Q. CD Ci <_) •!— ra X)
l/l l_ ra ra ^ 1 — ra
<t t3 Q. 3 OD -J
205
^ CO
^ cn
r- 1 —
1— CD r—
r- —
00 Ol ^ —
00
en
Ol 1— CO
cn
r— ^ cn •
— 1 —
— - 1 — 1 —
N
. ^ to
CM
<D
<— +->
1— CVJ to • +->
1
00
<_)
ra o
to 00 1— O)
to
cn
"Si-
c:
cn +-> to
t—
00
<u
+-> c
■t-J 1— <D C
+J
cn
s_
<D O
O) ^ -MO
CD
1 —
0)
(/>
a> o) l/l
3
<4-
OJ "O
CU <D l/l -O
CD
o
cu
05 C
cn CP 3 O c
cn
a;
q;
CD 3
cji cn o >> 3
cn
CD
i-
OJ E
CD CD 1- to E
CD
o
—1 <c
— 1 — 1 s <c
_i
a.
>>
to ^
to
O) O)
l-
c: Ol
+->
+-»
■t->
t.
i-
l-
i-
3
on C_>
c
ZJ
=)
3
2:
' —
o
o
o
O
o
1 — i-
<-)
u
u
s:
•r- OJ
+->
<D
cu
CD
ro
■M
+->
■M
+->
-t-" (J
U
o c:
O
_C
o
O -1—
_I
3
3
3
LL.
LI- Q-
XJ
E
<_>
c
o
•1—
•1— l/l
l/l
S-
Ci E
3 Olr-
uo CD
O O)
c c to
-M <->
+-> +->
3 •■- -H
c c
-o
O 1/1
o s- c
c=
l/l >1
TD CD
•1- i-
•1— CD l/l
+-> E
-a CD
3 •
CD +-> O
l/l to O)
(O M-
c zn
- S- C O
oix: >—
i_ <J-
3 a.
S- t/0
CJl +-> to O)
O •(-> O)
CP-i-
o
O) 1
C r—
■o
TO C
3 'J
•r- CD CD. +->
to -o c
o o
i- C l/l
C CD •<-
o o
l/l
r- (/O
o -t- -o a*
ro -H
•.- z
C l/l
•M o. c: s-
fO l/l
■H
O <D
•r- fO O
1 — 4_ O) l/l
3
•1- cn
i- cn
c « y-
ro ■)-> <J O)
4-> C
O) -1-
O l/l
O l/l C l/l
■— +->
ro to
3 -c
E c c: CD
•<- C <U to
o c
<_) J=
O O x:
Ol O 4- CP
a. to
o u
</i o
1_ .r- ..- +J
O E CD
Oj +J
•1- ■(-> +->
1 — CD y- I/)
Ol Q.
o
to to o
O T3 <^
e
+-> c
i- -o
t— s_ +->
O -r- O
O l/l
to
+J CD
cn CD +->
O) UJ "o +->
>— -o
CD
■o
C i- 3 i>o
X
to c
l/l i-
a> c
•I- c c
e i-H u_ CD
ro
S- CD
^ o
> i- o
O £3. O 3
C 1/1
to 3
r- CD •— -i-
S- -D
o
CD
uo
O +-> to -M
M- >>-a
>, O)
i_
O i-
c 3 CD u
l/l O 1 —
•r- O
cn CD
i- •
•1- C -r-
cn+J o Xi
1/1
on O l/l
■!-> -r- u-
C C CD to
O Q£
i- 1—
x:
<D U <D
>i c E •>-
•.- to s- E
CL O
o
u
+-> >
-O to TD
S- r- 3
E tJ
y-
l_
to -O (T3
3 r— +J O
Q. c — l/l
o z
TO
t- £= <U
+-> Ci C S •
CD to CD
<J 10
-o +->
0)
fO «—
l/l to "O
CD l/l (_) i-
CD C
>ir— CD
S- to -1- CX
-l->
J- to
(U
•<- -o <U
CD "O Q. • +->
+J cn s_
l/l O)
3 1—
l/l dJ JZ
l/l O >) C
o •
CD
l/l Q.
QJ +-> +->
(O O • S- CD
M- i- +-> c:
I. +-> •
to
x: <_)
O 3 >i -M E
O 3 l/l O
o -o
CD l/l
■(-> O) c
Ol l/l 3
O -1- -1-
M- E c
E to
C= Q.-I-
i_ . o •>- O
1/1 l/l +-»
O 3
cn
>. X
to l/l r— E O
■o s- o
1/1 <U 1/1
CD CD O <D T3
1/1 to to 1/1
CD y- i*-
£3. i-
>>-r— •!— ^
>> o
3
■t-> C CD
■O I/O O CD
1— E CD CJ.
■O >> CD
1— o
-o
O (O >
3 >> O S-
to O > E
3 ro i-
:C ^ <U
+->.£= -1- CD
c i- to o
-t-> 3 CD
o
•a
Q. +J r-
<t l/l C1X> 3
<: y- u
c/1 to 3
I/) to
c
c
<D
o
0) Q.
CJ
O
■(-> l/l
+->
1/1
•r- <
l/l
>>
X -C
CD
X
to
s_
3 -o
i-
o
O
(D CD
o
00
>— C • n3
<4-
1— c
CO
l/l
O •!- l/l
O •!- 1/1
o
(U
Q. Q.-0 OJ
ca. Q. -o
+■>
+->
Ol
CD •!- O
tO
CD -1-
l/l
E
o
CD
cn^ s-
cu
E to
J2
OJ
T3 O i-
s_
X3 O JD
S-
to o
fO
a.
O to >i Q.
o
o to >>
o
o .—
h-
00
-J "T) ^ (/)
CD
_J r-5 J=
_l
206
was pointed out by Torn et al. (1987) who suggested that liming practices would need to
become common if the present fertility of Alberta soils is to be maintained, primarily
to offset the effects of acidification caused by fertilization. This conclusion was
also substantiated by lurchenek et al . (1987) in their review of the potential for soil
acidification in Alberta.
In summary, Mayo (1987) stated that the effects of acidification documented for
Alberta, to date, require clarification to determine whether or not they are direct or
indirect effects and that a better understanding of forest processes is required to make
these clarifications.
9.3 ACIDIC DEPOSITION AND ALBERTA AGRICULTURE
Agricultural production contributed 10.2% of Alberta's gross domestic product
in 1981, with grain crops such as wheat and barley accounting for over 75% of Alberta's
farm cash receipts. Almost 30% of Alberta's land area is used for farming, with 12%
being cultivated at a given time (Alberta Agriculture 1982). Ecologically, agriculture
and grazing are dominant in four ecoregions of Alberta: Short Grass, Mixed Grass, Fescue
Grass, and Aspen Parkland, which in total cover about 25% of the province (Strong and
Leggat 1981). Thus, the effect of air pollutants on agriculture is of both ecological
and economic concern.
The Clean Air Act of Alberta sets maximum permissible levels of gaseous air
pollutants in the ambient air. No such standard is available for acidic precipitation.
There is, at present, no integrated network in Alberta for the monitoring of acidic
deposition (wet and dry) although independent, unrelated efforts for the measurement of
precipitation and its chemistry are in progress in the Province.
9.3.1 Wet Deposition Effects on Agriculture
Exposure to simulated acidic precipitation resulted in reduced yields in 14 out
of 19 agricultural species reviewed by Torn et al. (1987). However, there have been no
field surveys documenting a reduction in yield due to ambient acidic deposition levels
in the Province.
There is little evidence for a linear dose-response function for simulated
acidic precipitation and plant injury. However, below pH 3.5, dose -response does
approach linearity, with a yield loss of approximately 5% per decrease of one pH unit.
The formation, development, and survival of pods, flowers, and fruits are
sensitive to simulated acidic rain at moderately low pH values (below 4.0). Foliar
injury resulting from exposure to simulated acidic deposition can lower marketable yield
of truck crops, lower plant resistance to pathogens, and has been linked with reduced
plant productivity.
The threshold for simulated acid rain-induced foliar injury to agricultural
crops was found to be between pH 3.0 and 3.5 for 36 crop species reviewed by Torn et al.
(1987). In decreasing order (most to least) of sensitivity to simulated acidic
precipitation, these crop types were: root, leafy, cole, legume, fruit, grain, and
leafy and seed forage crops. The potential for economic loss was highest in leafy,
cole, and fruit crops. Monocots such as wheat, barley, and timothy were found to be
resistant to foliar injury from simulated acid rain above a pH of 2.5. At current
207
ambient levels of precipitation acidity there is little risk of foliar injury within the
Province of Alberta; however, increased emissions may pose a risk to sensitive plants in
the future.
Under the present conditions, it is unlikely that S or N in acidic deposition
could be a significant source of foliar fertilizer to crops, or pose a risk of salt
damage to crops in Alberta at current levels.
In summary, at current levels of precipitation acidity in Alberta, acidic wet
deposition is not a concern for agriculture at this time.
9.3.2 Dry Deposition Effects on Agriculture
If the air quality standards for sulphur and nitrogen dioxide are met in the
Province of Alberta, there should be no adverse effects on agriculture as a result of
dry deposition under most conditions and with most species of agricultural plants.
However, some studies reviewed by Torn et al . (1987) suggest that if the most sensitive
agricultural species are exposed to sulphur dioxide at concentrations slightly higher
than permissible standards under conditions conducive to gas exchange, they may be
injured. To injure the most sensitive species, average SO2 concentrations of 0.05 to
0.5 ppm for several hours are usually required. It is important to note that pollutant
peaks are included within the average values. Thus, the avarage values do not represent
the true exposure. The effects of ambient sulphur dioxide on the yield of agricultural
crops are shown in Table 39.
Nitrogen dioxide concentrations of 0.25 to 0.50 ppm for long periods of time
are generally required to induce injury in sensitive agricultural plants. However, a
few studies have indicated injury to plants at concentrations at or below the maximum
permissible Alberta concentrations. For this reason, Torn et al . (1987) believe that if
the provincial standards are adhered to for acute nitrogen dioxide exposures, no injury
should occur in agricultural species. A few experiments reviewed by Torn et al . (1987)
did show plant injury under chronic exposure conditions at or above permissible concen-
tration maxima. These authors questioned whether the standards for chronic exposure were
sufficient to provide protection for sensitive agricultural species. The susceptibility
or sensitivity to nitrogen dioxide exposure of a number of agricultural crop species
commonly grown in Alberta is shown in Table 40.
Exposure to ozone concentrations of 0.03 ppm (for very sensitive species) and
0.10 ppm (for plants with intermediate sensitivity) for several hours is required to
induce injury in agricultural plants. Research indicates that the maximum permissible
concentrations for ozone specified under the Alberta Clean Air Act are sufficient to
protect agricultural plants in the Province (Torn et al. 1987). Injury to agricultural
species of intermediate sensitivity due to chronic exposures is generally not seen at
long-term average ozone concentrations below 0.05 ppm (Torn et al . 1987). Agricultural
crops grown in Alberta which are known to be relatively sensitive to ozone are shown in
Table 41 .
A review of the most recent research on the effects of hydrogen sulphide on
agricultural crops indicates that present provincial standards are sufficient to protect
plants (Torn et al . 1987) .
208
Table 39. Effects of ambient sulphur dioxide on yield of various
agricultural species.
Crop & Harvest Characteristics Percentage of Control Value
Spring Canola
(yield) 90.9
Alfalfa
(yield) 81.0
Oats
(yield) 76.1
Spring Wheat
(yield) 73.4
Red Clover
(yield) 63.6
Winter Rye
(yield) 57.7
Winter Wheat
(yield) 55.6
Exposure time: 4.3% of monitoring time
Concentration: 0.44 ppm - during exposure time*
0.083 ppm - average for monitoring time**
* The exposure time was calculated by summing all time intervals,
At = 10 minutes, with a mean SO2 concentration greater than
or equal to 0.10 ppm.
** The monitoring time is essentially equal to the exposure time
of the test plants.
Source: Guderian and Stratmann (1968)
Table 40. Susceptibility of various agricultural species which
occur in Alberta to nitrogen dioxide.
Plant Species Susceptible Intermediate Resistant
Alfalfa
+
Annual bluegrass
+
Barley
Kentucky bluegrass
+
Oats
+
Potato
+
Red clover
+
Rye
+
Sweet corn
+
Wheat
+
Asparagus
+
Cabbage
+
Carrot^
+
Celery^
+
Kohl rabi
+
Leek
+
Lettuce
Onion
+
Tomato
+
^Different investigators reported different degrees of
susceptibi 1 ity
Adapted from the original table in Legge et al. (1980)
Source: National Academy of Sciences, U.S. (1977b)
210
Table 41. Agricultural crops grown in Alberta which are known to
be relatively sensitive to ozone.
Alfalfa (Medicaqo sati va)
Barley (Hordeum vulgare)
Bean ( Phaseolus vulgaris)
Red clover (Trifolium pratense)
Corn, sweet (Zea mays)
Grass, bent ( Agrostis palustris)
Grass, brome (Bromus inermis)
Grass, crab ( Digitaria sanguinalis)
Grass, orchard (Dactylis glomerata)
Muskmelon (Cucumis melo)
Oat ( Avena sati va)
Onion (Al lium cepa)
Potato (Solanum tuberosum)
Radish (Raphanus sativus)
Rye (Secale cereale)
Spinach (Spinacea oleracea)
Tomato ( Lycopersicon esculentum)
Wheat (Triticum Aesti vum)
Source: Hill et al . (1970)
211
Because injury to sensitive agricultural species has been observed during
chronic exposures at or near maximum permissible levels of certain gaseous pollutants
when present singly, there is concern over the possible more than additive effects of
these pollutants at the same concentrations when present in combination.
More than additive, additive, and less than additive effects resulting in
decreases in growth and yield have been observed with exposures to mixtures of SO2 and
03, SO2 and NO2, and NO2 and O3. In addition, researchers have found that in nearly every
instance, exposure to a mixture of SO2, NO2, and Oa causes a greater loss in plant
growth and yield than the exposure to single gases or to mixtures of two gases. Growth
and yield responses to nitrogen dioxide in pollutant mixtures occur in the nitrogen
dioxide concentration range of 0.05 to 0.30 ppm. This is well below the current air
quality standard and within the ambient concentration range of NO2 within Alberta.
The decrease in growth and yield caused by nitrogen dioxide in the presence of sulphur
dioxide and/or ozone ranges from 5 to 20% at concentrations of nitrogen dioxide that
cause little or no injury when the pollutant is present singly.
9.4. ALBERTA SOILS SENSITIVE TO ACIDIC DEPOSITION
Turchenek et al . (1987) provide a detailed overview of the soils of Alberta
including their classification and the amounts of each major class in the province
(Table 42), processes involved in soil formation and a general description of the major
orders of the soils found in the province.
The sensitivity to acidic deposition of the various major soil classes is
discussed at length by Turchenek et al. (1987) under three different approaches: soil
sensitivity and mapping; qualitative descriptions of possible soil responses to
acidification; and modelling approaches using dose-response methodology. Only the
predictive model by Bloom and Grigal (1985) is described here; for other modelling
approaches refer to Turchenek et al. (1987). The Bloom and Grigal (1985) model bases
its predictions on two levels of acid input, (1) low - 0.1 Kmol (H^) ha ^ y ^
and (2) high - 1.0 Kmol (H^) ha"^ y~\
9.4.1 Soil Sensitivity and Mapping
Sensitivity refers to the ease with which soils can be affected or influenced
by acidic deposition. Schemes for rating sensitivity of soils to wet and dry acidic
deposition have been developed for the purposes of grouping geographic areas into
sensitivity classes. The methods and parameters used to determine sensitivity vary
according to the author and geographic region for which it was developed. In Alberta,
Holowaychuk and Lindsay (1982) have developed such a system which was used to classify
soil sensitivity in the Sand River area of northeastern Alberta (near Cold Lake). None
of the systems in current use consider deposition impacts other than on pH and the
exchange complex. Effects on organic matter turnover and on the dynamics of major
nutrients such as N, P, and S are not considered. Insufficient data exist at the present
time to incorporate impacts on organic matter and nutrients into the sensitivity clas-
sification systems in use. Therefore, these sensitivity ratings provide a first
approximation of the potential impacts on some soil properties, and have been developed
to identify specific soil types sensitive to acidic deposition and their locations.
212
Table 42. Areas of the soil orders in Alberta.^
Soil Order Area (km^ x 1000)
Chernozemic 141.5
Solonetzic 43.0
Luvisolic 203.1
BrunisoUc and Podzolic 52.9
Regosolic 7.4
Gleysolic 21.6
Organic 104.6
Cryosolic 43.9
Nonsoil " Rockland, Icefields, 26.4
- Freshwater 16.8
Total 661.2
^Adapted from Holowaychuk and Fessenden (1987)
213
Holowaychuk and Fessenden (1986) have classified and mapped the soils of Alberta
with respect to their sensitivity to acidic deposition. Soils of the province were
described by delineating major soil landscape units and indicating the properties of
both dominant and subdominant soils in each map unit. Mapping is on a broad regional
scale resulting in inclusions and variation within soil groupings to be ignored.
Attributes of soils used to differentiate map units include taxonomic class, pH, texture,
and type of parent material.
9.4.2 Chernozemic Soil Impacts
Holowaychuk and Fessenden (1987) have identified the following Chernozemic soil
types as being sensitive to acidic deposition: sandy, and some coarse loamy, Orthic
Brown, Rego Brown, Orthic Dark Brown, Rego Dark Brown, Orthic Black, Rego Black, and
Eluviated Black. In addition, a small area of Dark Brown soils in the Cypress Hills was
indicated as being moderately sensitive. Summary statistics on Chernozemic soils in
Alberta potentially sensitive to acidic deposition are presented in Table 43. The
Chenozemic soils which appeared to be highly or moderately sensitive to acidic depo-
sition are all slightly acidic to start with and are subgroups found on weakly to
moderately calcareous parent materials.
Application of the Bloom and Grigal (1985) model to a few Alberta Chernozemic
soils indicated that response to addition of about 0.1 kmol (H^)ha ^ y ^ and
1.6 kg S ha ^ y ^ would be slight. Severe effects such as decline in pH and base satura-
tion to the point of inducing aluminum toxicity would occur after about 200 years. A
loading of 3 kmol (H^)ha ^ y ^ and 48 kg S ha ^ y ^ would have drastic effects within
50 years in some soils of this group. Sandy soils of low CEC were found to be the most
susceptible to acidification. Adverse effects were predicted to occur soonest in those
soils which already have relatively low pH and base saturation levels.
An overriding issue in considering the effects of acidic deposition on Cherno-
zemic soils is that most areas with such soils are under cultivation and are fertilized.
The strong acidifying effects of fertilization are well documented and have been dis-
cussed elsewhere in this synthesis. In several reviews of this problem in recent years,
it has been concluded that liming should become a general practice in order to restore
productivity in lands affected by acidification due to fertilization practices. Any
acidification caused by atmospheric deposition will thus be neutralized as well. Where
liming practices are, in effect, soil responses to acidic deposition are not likely to
be an important concern.
9.4.3 Solonetzic Soil Impacts
The area of Solonetzic soils in Alberta is about 42,960 km^. None of these
soils are presently regarded as being highly sensitive to acidic deposition, but about
one-third are considered to be moderately sensitive (Holowaychuk and Fessenden 1987).
Those which are relatively sensitive to acidic deposition are soils of the Solonetz
great group which occurs on coarse parent materials or on shallow glacial deposits
overlying residual materials. The Gray Solonetz subgroup is also in the moderately
sensitive category because it lacks the Chernozemic type A horizons characteristic of
prairie soils. The A horizon allows more leaching and lower colloid and base cation
content in this horizon, thus lowering buffering capacity.
214
Table 43. Chernozemic soils sensitive to acidic deposition.
Area
(km')
% Sensitive
High
Moderate
Brown Chernozemic
Dark Brown
Black
Dark Grey
Sandy/Coarse Loam
34,306
40,418
2641
6848
N.E.
16%
14%
4%
52%
All
<1%
N.E. = No estimate
Adapted from Turchenek et al . (1987)
215
The Bloom and Grigal (1985) model was applied to two series of Solonetzic soils with low
pH values (Camrose 4.8, Brownfield 4.9, measured as calcium chloride) using an acid
input level of 0.1 kmol(H^) ha ^ y ^ . At this low acid input level, the model predicted
that for the tested soils series, only a minor lowering of pH and base saturation and
increases in aluminum content would occur. The pH drop in 100 years was predicted to be
approximately 0.2 units for the Camrose soil and 0.4 units for the Brownfield soil under
high acid input levels of 1.0 KmolCH"*") ha ^ y ^; base saturation changes were also
slight with Brownfield soils exhibiting the most decrease. The aluminum content was
predicted to reach toxic levels in 250 years for Camrose soils and in 350 years for
Brownfield soils.
Turchenek et al . (1987) also pointed out that although coarse Solonetzic soils
were not addressed in the modelling, responses in pH, base saturation and aluminum
content to acidic deposition would occur at a faster rate than the evaluated soils
because of their texture. They predicted aluminum toxicity as a result of acid
deposition would occur within 100 years.
9.4.4 Luvisolic Soil Impacts
The area of Luvisolic soils in Alberta is approximately 203,000 km^. Most
of these soils are considered to be moderately sensitive to acidic deposition with less
than 5% being considered sensitive. Sensitivity is caused by a relatively low base
status in the A horizon of Luvisolic soils. In addition, these types of soils are
formed under forest vegetation. These soils, because of the inefficiency of nutrient
cycling, may have low base saturation as well as low total exchange capacity.
The responses of five series of Luvisolic soils were evaluated using the Bloom
and Grigal (1985) model. These were: Culp, Leith, Breton, Nosehill, and Tom Hill. The
Tom Hill soil was a Podzolic Gray Luvisol and the Leith was a coarse Dark Gray Luvisol.
All other test soils were Orthic Gray Luvisols which varied in texture, with the Breton
and Nosehill being of the fine loamy variety and the Culp being coarse textured.
The modelling results for all test soils are shown in Table 44 for both high
and low acid inputs. These predictions show pH depressions on all sub-groups tested at
high acid inputs after 100 years. The most substantial changes occurred in the coarse
textured Culp and the fine-loamy Tom Hill sub-groups. The model also predicted a small
drop in base saturation in all sub-groups at low acid inputs. Base saturation at high
acid inputs decreased, following similar trends to those detected for pH. For example,
in the Culp sub-group, the model predicted base saturation to decrease from 87% to 44%
within 100 years at the high acid input, while the least affected sub-group (Leith)
indicated base saturation decreased from 89% to 70% within 100 years. Because of the
higher clay and organic content of this latter sub-group, it has a higher CEC and
exchangeable base content than the soils of the Culp series. Over the course of time
all soils in the tested series were predicted to lose almost all of their base cations.
Aluminum content in the soils was predicted to increase only slightly at low
acid inputs. Nosehill soils, the most acidic to start with, started with fairly high
levels of aluminum (3 yM) which were predicted to rise to 9 yM within 100 years at
high acid input. The model predicted increases in aluminum content to toxic levels for
all soils over time. In a decreasing order of susceptibility to this predicted change
216
Table 44. Modelled predictions (Bloom and Grigal 1985) for soil pH
responses to acid inputs.
Acid
Time
SUD-
Sensi-
Initial
Input
Projec-
Frame
Soil Type
Texture
Group
ti vity
pH
Level
ted
pH
(Years)
LUVISOLIC
Moderate-
High
nPTHTr ARAY
UfxiniL. UfxHI
Till n
H i nh
n 1 y 1 1
J . o
1 0
1 . u
4.
9
1 on
DARK GRAY
Coarse
Leith
Moderate
5.9
1 .0
, D
100
nPTHTP fiRAY
p-i np— 1 n;i mv/
r 1 1 1 c 1 vjaiiiy
R rp 1" n n
Hi ah
n 1 y 1 1
S ft
J . u
1 0
1 . \J
5.
,3
1 no
PODZOLIC
Fine-loamy
Nosehi 1 1
High
4.4
1 .0
4.
,0
100
DRTHTr HRAY
Ft ri-D — 1 n3m\/
r 1 1 It; 1 uaiiiy
Tom Hill
1 uiii mill
Hi nh
n 1 y 1 1
J . H
4.
,7
1 00
BRUNISOLIC
ELUVIATED
DYSTRIC
Sandy
Fi rebag
High
5.1
0.1
4
.6
100
5 . 1
1 . 0
3
.7
50
ELUVIATED
DYSTRIC
Fine-loamy
Robb
High
4.2
1 .0
3
.7
50
4.2
1 .0
3
.0
100
ORGANIC
High
NT
ORGANIC
CYROSOLIC
Low
NT
GLEYSOLIC
Low
NT
REGOSOLIC
Low
NT
Acid input: High level 1.0=1.0 kmol (H+) ha'^ y-^
Low level 0.1 = 0.1 kmol (H+) ha"^ y"^
NT = Not tested
Adapted from: Turchenek et al . (1987)
217
in the aluminum concentration, the soils were: Nosehill, Gulp, Tom Hill, Breton, and
Leith.
The responses of Luvisolic soils to acidic inputs are generally similar to
those of Chernozemic and Solonetzic soils. The response rates are highest in coarse
textured soils with low CEC and exchangeable base content. The pH and base saturation
in Luvisolic soils are relatively low. Therefore, critical levels of pH and aluminum
content may be reached within shorter periods of time under acidic deposition than they
would be in other soils.
9.4.5 Brunisolic and Podzolic Soil Impacts
About 40% to 45% of the 53,000 km^ of Brunisolic soils in the province are
considered to be highly sensitive to acidic deposition (Holowaychuk and Fessenden 1987).
These soils are mainly Eluviated Eutric Brunisols and Eluviated Dystric Brunisols
developed on glaciof 1 uvial and eolian deposits. Other types of Brunisols in the Province
have a low sensitivity to acidification. A small area of about 390 km^ of Orthic
Humic Podzols and associated Dystric Brunisols occurring west of Grande Cache are also
rated as being highly sensitive to acidic deposition (Holowaychuk and Fessenden 1987).
The modelled predictions (Bloom and Grigal 1985) for two sensitive soils of
this group are shown in Table 44 for both high and low acid inputs. The Firebag soils
were predicted to have a strong response to low level acid inputs with effects beginning
to show within 100 years. This response was the highest for all the soils tested using
the Bloom and Grigal (1985) model. Along with the observed pH depression, the base
saturation of Firebag soils dropped from 31 to 10% and aluminum content increased from
0.9 to 2 yM at low acid inputs. Both soils tested were predicted to react drastically
to high acid inputs with pH and base saturation dropping substantially and aluminum
content rising within 50 years. The calculated aluminum content for both soils was
predicted to be over 100 yM within 50 years, an alarmingly toxic level. Although the
initial aluminum content of these soils is not known, it is suspected to be in the order
of 10 yM for the Robb soil and less than that concentration for the Firebag soils.
Therefore, the predicted rise is dramatic.
The acid deposition response simulations indicate that sandy Brunisolic soils
are among the most sensitive in Alberta and that particular attention should be given to
these soils and the ecosystems they are a'ssociated with. Finer textured Brunisols are
considered to be less sensitive, but the simulations indicate that there could be
problems related to high aluminum in soils of very low pH. It should be noted, however,
that although high aluminum contents were predicted for the sandy Firebag soils, the
lack of easily weatherable minerals in these soils may not allow the system to contribute
the high aluminum levels predicted.
9.4.6 Organic and Organic Cryosolic Soils
Almost all of Alberta's 104,500 km^ of Organic soils are considered to be
highly sensitive to acidic deposition (Holowaychuk and Fessenden 1987). Organic soils
throughout the Province are generally in the very strong to medium range of response to
acidity, have low base saturation, and on a volume basis, low base status. If subjected
to further acid inputs, the pH would be reduced and further depletion of base saturation
would occur.
218
The 43,000 km^ of Organic Cryosols in the Province are rated as having low
sensitivity to acidic deposition (Holowaychuk and Fessenden 1987). Most Organic Cryosols
are highly acid in reaction and it is generally conceded that additional acid will not
cause further acidification or alter their base status.
The soil response simulation model was not run for either Organic or Organic
Cryosol soils. The Bloom and Grigal model was developed for use on mineral soils and
was not considered applicable to organic soils because model input parameters such as
pH-base saturation relationships, and aluminum activity coefficients, differ between
organic soils and mineral soils (Turchenek et al. 1987).
9.4.7 Gleysolic Soil Impacts
There are about 21,000 km^ of Gleysolic soils in Alberta. These have been
rated as having low sensitivity to acidic deposition (Holowaychuk and Fessenden 1986).
A small 836 km^ area northeast of Grande Cache is the only exception, and is rated
as having a high sensitivity. No acidity response simulation was run for this type of
soil because of its lack of sensitivity (Turchenek et al. 1987).
9.4.8 Reqosolic Soil Impacts
There are approximately 7400 km^ of Regosolic soils in Alberta. They are
not considered to be sensitive to acidic deposition because they receive continuous
replenishments of calcareous materials by means of alluvial sedimentation. Most of
these soils are found in the Peace-Athabasca Delta. No acidity response simulation was
run for this soil type because of the suggested lack of sensitivity to acidification.
9.4.9 Impacts on Rocklands and Rough-Broken Lands
Rockland and Rough-Broken Lands account for about 35,000 km^ of Alberta's
land area. Two groups of these lands are considered to be sensitive to acidic deposi-
tion, the Precambrian Shield area of northeastern Alberta and Cordilleran areas of almost
barren clastic sedimentary rock and mineral materials. Coarse textured Brunisolic and
Podzolic soils occur within the Rockland areas.
9.5 EFFECTS OF ACIDIC DEPOSITION ON SOIL MICROORGANISMS AND PROCESSES
The majority of studies dealing with the impact of acidic deposition on quali-
tative and quantitative aspects of the soil microbial community have dealt with chronic
effects in the field and, therefore, the results may be applicable to some Alberta
soils. It can be concluded that acidification of a naturally acidic forest soil (in
Alberta, the pH of surface soils in many coniferous forests ranges from 4.6 to 6.0) to
approximately pH 3.0 or less, results in a significant reduction in total microbial
biomass. Bacteria are adversely affected at pH 4.0, and fungi appear to be less
sensitive to acidification.
Factors in the N cycle such as ammonif ication, nitrification, nitrogen fixation,
and symbiotic relationships are especially sensitive to acidity and could be adversely
affected by acidic deposition. There is little reduction in the rates of ammonif ication
until the pH reaches 3 or less but nitrification and nitrogen fixation are inhibited in
soils below pH 6. It appears clear that the 1 egume-Rhi zobi um symbiosis is acid
sensitive.
219
Although studies concerning the relationship of acidic deposition and
mycorrhizae are few in number, they indicate that these symbiotic relationships are
largely resistant to chronic exposures. Most ectomycorrhi zal plants are associated with
a diverse array of acid tolerant fungi which probably provide a strong environmental
cushioning capacity against chemical change. In possible contrast, at least some VA
mycorrhizal fungi are acid sensitive. Reductions of pH by 0.5 to 1 unit may render
certain of these species nonfunctional with consequential reduction in plant growth.
Most of the laboratory simulation studies reviewed by Visser et al. (1987) used
extremely low pH levels that did not occur in ambient conditions. Because of this,
these studies are not considered relevant to the provincial problem. However, based on
a review of both laboratory and field studies by these same authors, it is clear that a
reduction in the pH of naturally acidic forest soils to 3.0 or less will have an inhibi-
tory effect on soil respiration. However, the high buffering capacity of litter and
decaying plant residues and the probable presence of microbial flora adapted to acidity
in such situations makes it likely that extremely high dosages of acidic rain or sulphur
dioxide would be necessary to reduce microbial respiration to any substantial degree.
It is not known if this would also be the case in agricultural grasslands where less
acid tolerant microflora may reside.
Controlled laboratory and field experiments have also shown that simulated
acidic rain of pH 2.0 or fumigations with sulphur dioxide up to 530 ppb are necessary to
inhibit litter decomposition. It is unlikely that either situation will ever occur in
Alberta under present conditions or if current environmental standards are maintained.
9.6 SULPHUR MICROBIOLOGY IN THE ALBERTA CONTEXT
The review of soil sulphur microbiology prepared by Laishley and Bryant (1987)
is of particular relevance to Alberta. The topic deals with elemental sulphur breakdown
and bears direct relevance to soils and soil acidity because of the importance of the
sulphur industry in the Province and the storage of sulphur and fugitive sulphur dust
problems associated with it.
Only key highlights of the Laishley and Bryant (1987) report will be presented
here:
1. The sulphur oxidizing colourless bacteria and particularly the thiobacilli
are the principal microorganisms responsible for the breakdown of elemental
sulphur in Alberta.
2. Biological oxidation of inorganic sulphur can create acid soils, with
accompanying leaching of nutrients such as Fe^^ and Al^^, resulting in tox-
icity to plants. Bacterial acid production can also cause direct injury
to plants.
3. Different microorganisms can oxidize sulphur at different rates. The most
prolific oxidizers belong to the genus Thiobaci 1 lus . different species of
which oxidize sequentially as acid conditions in the soil change under
oxidizing conditions created by the bacteria themselves (Laishley and
Bryant 1987).
Factors affecting the oxidation of fugitive (wind blown) sulphur include:
a. the particle size and microcrystal 1 ine structure of the sulphur
itself;
b. the types of sulphur oxidizing bacteria present at the deposition
site. For example, the acidophilic thiobaccilli would not be active
in soils with basic pH. Conversely, in more acidic environments, it
would be unusual to find less acidophilic bacteria playing a dominant
role. Oxidation processes can also be limited by the formation of
biofilms of bacteria on the upper molecular layer of sulphur parti-
cles (Bryant et al. 1983). These films, formed of bacterial cells
and colonies and their attachment glycocalyx, effectively seal off
the sulphur to continued oxidation (Takakawa et al. 1979; Costerton
and Irvin 1981; Ladd 1982; and Bryant et al. 1983);
c. the soil environment. Bryant et al. (1985) have shown that bacterial
oxidation process decreases with temperatures above or below the
optimum of 28°C. They also found that temperatures of 5°C or 37°C
effectively stopped sulphur oxidation by Thiobaci 1 lus albertis . The
bacteria were reactivated if temperatures were increased above 5°C but
died at 37°. This information is important to the Alberta situation
because it suggests that to minimize sulphur oxidation, the sulphur
should be kept cool. Alberta is noted for its long winters and cool
overall climate which should assist in reducing oxidation rates. All
bacteria known to oxidize sulphur in Alberta are capable of with-
standing freezing, and reactivating when warmer conditions return
(Laishley and Bryant 1987).
Soil type, especially pH, texture, and base saturation, has also
been shown experimentally to affect bacterial oxidation of sulphur
(Laishley and Bryant 1987). The percentages of soils in Alberta with
pH less than 6.0, based on soil testing data, are shown in Figure 11.
The third soil factor determining the rate of sulphur oxidation
by bacteria is the moisture and nutrient status of the soil. Thio-
bacilli are aerobic and will not grow in waterlogged soils (Laishley
and Bryant 1987). Laishley and Bryant (1985) have also shown that
under dry soil conditions, sulphur oxidizing activity was low. What
is required for bacterial oxidation is a level of moisture close to,
but not exceeding, soil moisture holding capacity. Microorganisms
require the same nutrients as plants for successful growth. Sulphur
oxidation, in fact, may be enhanced by N-P fertilization (Bloomfield
1967) which suggests that agricultural soils in Alberta associated
with gas plants may have high sulphur oxidation.
Fugitive (wind blown) sulphur from existing stock piles or sulphur opera-
tions will create an ideal soil environment for the sulphur oxidizing
Thiobacilli, with the end result being acidification of the soil. Long
term studies of SO2 contamination of soils near a gas plant source did
not show trends in acidification (Lore 1984). This finding suggests that
221
Figure 11. Location of soil testing areas in Alberta, and the
percentage of cultivated soil with a pH of 6.0 or
less for each area. (Taken from Penney et al . 1977)
222
acidification of soils from sulphur dioxide deposition is not as severe a
problem in Alberta as wind blown elemental sulphur dust.
Liming of acid soils polluted with sulphur only masks an existing
problem and really provides a more favourable environment for continued
microbial sulphur oxidation and acidification (Laishley and Bryant 1987).
Reports by Nyborg and Hoyt (1978) and Ivarson (1977) indicate that liming
causes a temporary increase in organic nitrogen mineralization but also
results in an increase in microbial activity. Increasing the nutrient
status of such soils while leaving the elemental sulphur as a microbial
substrate will tend to favour continued oxidation by thiobacilli and,
hence, renewed acidification. This means that the original reason for the
liming will reoccur quickly unless the sulphur is removed.
6. Decommissioning of sour gas plants and recovery of bulk sulphur storage
blocks are becoming more commonplace in Alberta. Unfortunately, many of
the sulphur blocks were laid directly on bare soil and it is estimated
that there may be 20% to 30% of the sulphur left in the soil at the end of
the cleanup process (Hyne, pers. comm. in Laishley and Bryant 1987). These
soils will be subject to intensified microbial acidification processes,
particularly if the remaining sulphur is in powdered form.
9.7 SURFACE WATtR ACIDIFICATION STUDItS IN ALBERTA
Hesslein (1979) surveyed 20 lakes in the Alberta Oil Sands Environmental
Research Program (AOSERP) study area near Fort McMurray in order to determine their
susceptibility to pH change resulting from acidic deposition. The lakes were sampled
for only a short period of time; October 6 - 1 0, 1976. Most lakes in the region were
found to have high pH values and high alkalinity. However, lakes could be divided into
two groups: (a) pH and alkalinity in the range of 6.18 - 7.49 pH units and 342 -
811 yeq L ^, respectively, and (2) pH and alkalinity in the range of 7.89 - 8.32 pH units
and 966 - 3090 yeq L''^, respectively. The first group of lakes is located at the north-
west corner of the study area, whereas the second group of lakes is located in the more
a''kaline region along the Athabasca River. Hesslein ( 1 979) suggested that the first
group of lakes might be susceptible to serious pH alterations if the average pH of rain
is below 4.0. However, this is unlikely to happen. Sulphur emissions in Alberta are
controlled by provincial legislation, allowing less than one percent of the total oxides
of sulphur to be emitted into the atmosphere. The Sudbury, Ontario area, with much
greater emissions of sulphur dioxide than Alberta, has an average precipitation pH of
4.0 to 4.5. Thus, it is unlikely that precipitation pH of less than 4.0 could occur in
Alberta (Hesslein 1979) .
Erickson and Trew (1987) compiled historical water chemistry data for 875 lakes
throughout Alberta. In addition, 107 lakes from the northern part of the Province were
surveyed and water quality data were collected. The study was designed to identify the
sensitivity of lakes to acidification. Based on earlier Canadian studies, the following
parameters were measured or compared in each lake: pH, calcium, and alkalinity. The
results of the study are shown in Table 45. Based on these findings, Erikson and Trew
(1987) suggested that lakes in widely diverse areas of the province are potentially
223
Table 45. Indicator parameters used to classify the sensitivity to
acidification of Alberta lakes.
Parameter
Range
% of Surveyed
Range Sensitive
Lakes in a
to Acidification
PH
3.4 - 10.6
17
Calcium (mg L~^)
0.1 - 8.0
18,
.6
Alkalinity (mg L~^)
Undetected
- 7772 9
.7
Adapted from: Erickson and Trew (1987)
224
sensitive to the effects of acidic deposition. A large number of sensitive lakes are
located on the Canadian Shield in the northeastern part of the Province. Several other
sensitive lakes were detected in the Rocky Mountains and in particular, Jasper National
Park. The northern upland areas of Clear Hills, Swan Hills, Caribou and Birch Mountains
also contain lakes which, based on the survey results, indicate potential sensitivity to
acidic deposition. A number of these lakes appeared to be naturally acidic based on
their colour and muskeg drainage characteristics.
In Alberta, streams and rivers have been routinely sampled by Environment
Canada and Alberta Environment since 1969. Very few lakes, particularly the ones at
high altitudes, have been sampled by either agency. In general, most Alberta rivers and
streams have high pH values (above 7.5) and high alkalinity (above 1000 yeq L ^)
(Environment Canada 1982).
Most of Alberta, except the northeast corner of the province, is underlaid by
carbonate rich minerals (limestone, dolomite, slightly calcareous rocks) with medium to
high buffering capacity. The northeast corner of the province is in the Precambrian
Canadian Shield and is underlaid by granitic gneisses and quartz sandstone with low
buffering capacity. Shewchuk (1981) conducted regional surveys of rain, snow, and lake
water chemistry in several locations on and near the Precambrian Shield of western
Canada. The areas surveyed included part of Saskatchewan (east to northwest section),
part of Manitoba, The Northwest Territories, and the northeastern section of Alberta.
Snow and rain in the region exhibited average pH values of 5.0 and 6.4, respectively,
indicating that the area was not receiving significant amounts of acidic deposition.
The pH of lakes on the Precambrian Shield averaged 7.3; the pH of lakes on the fringe or
off the Precambrian Shield averaged 8.1. Sulphate, alkalinity (as CaCOa), and calcium
ion concentrations in the Precambrian Shield averaged 2.7, 13, and 4.3 mg L ^,
whereas, in the fringe or off-shield lakes the average was 1 5, 200, and 55 mg L~^,
respectively. Although Precambrian Shield lakes are highly sensitive to acidic deposi-
tion due to lack of buffering capacity, there is little evidence to suggest that at
present they are undergoing significant acidification. Acidification studies conducted
on lakes located in the Precambrian Shield of Nova Scotia, Ontario, and Quebec, however,
suggest that these lakes have become acidic over the last few decades. Acidic deposition
is reported to be the cause of the lake acidification. Although the mechanism of their
acidification is a subject of much discussion, it appears that lakes located in this
sensitive Shield area are susceptible to acidification. Therefore, one can make a
cautious prediction that lakes located in the northeast corner of Alberta may also be
susceptible to acidification.
Three reports coordinated by Alberta Environment and Western Canada LRTAP (Long
Range Transport of Air Pollutants) concerning the acidification potential for northern
Alberta are nearing completion. They include: an inventory of freshwater systems and
their sensitivity ratings; a mapping of the northern soils and hydrogeol ogy ; and, a
statement of target loading for northern Alberta (David Trew, Alberta Environment,
personal communication cited in Telang 1987). When completed, they will provide a
valuable tool for examining future scenarios of, and possible mitigative measures for,
aquatic acidification in northern Alberta, particularly in the context of oil sands
production and development.
225
9.8 ALBERTA HYDROGEOLOGY AND GEOLOGY
In Alberta, groundwaters are characterized by relatively high pH values (6.5 to
8.0) and high acid neutralization capacities. Although groundwater chemistry data in
the northern part of the Province are limited, Schwartz (1979, 1980) suggested that even
in muskeg areas, groundwater is characterized by pH values in the 6.5 to 8.0 range.
Much of Alberta is covered with relatively thick deposits of glacial till that
have high acid neutralization capacity. It is doubtful if acidification of groundwater
in these deposits would occur in the short term given the present level of atmospheric
loading. It may, however, be possible to see the residues or byproducts of the acid
neutralization process in selected aquifers.
Krouse et al . (1984) have found evidence in the West Whitecourt area to suggest
that the products of acid neutralization reactions have already begun to reach the water
table.
Aquifers in Alberta which are most likely to show some documentable response to
acidic deposition would be characterized by:
1. low buffering capacity;
2. rapid recharge;
3. shallow, short flow regime; and
4. high permeability.
Geological settings that are likely to reflect the above-mentioned hydrogeol ogi cal
characteristics would occur where aeolian or glaciof luvial deposits overlie dense low
permeability till or bedrock.
Groundwater resources in Alberta are utilized extensively for domestic,
agricultural, municipal, and industrial purposes. Changes in groundwater as a response
to acidification could have a dramatic effect on the economy and public health of the
Province. Experiences in other parts of the world, such as Sweden, coupled with the
evidence cited from Whitecourt (Krouse et al . 1984), should alert Albertans that even in
our well protected environment, the potential exists for acidity-induced effects to be
felt in our groundwaters.
9.8.1 Geological Information Bases for Alberta
9.8.1.1 Bedrock Geology. As part of his review of the geological and hydrogeological
aspects of Alberta pertaining to acidic deposition, Campbell (1987) compiled a
1:1,000,000 scale bedrock geology map of the Province. This map was based on information
contained in maps compiled by Green (1972). Included in the Campbell map sheet are data
covering the area and types of bedrock, and metal formations within the Province. These
data are essential for the determination of bedrock and groundwater sensitivites to
acidic deposition.
9.8.1.2 Surficia! Geology. Campbell (1987) has also compiled a 1:1,000,000 scale
surficial geology map of Alberta. This map was based on information assembled from the
following sources: Geology Branch of the Research Council of Alberta, Geological Survey
226
of Canada, the Exploratory and Reconnaissance Soil Survey of Alberta Research Council,
and Special Geomorphic Landforms Maps (Environment Canada and Alberta Energy and Natural
Resources, Parks Canada Biophysical Inventory, and the Atlas of Alberta 1969). This
inventory is critical because of the effects that surficial materials have on shallow
groundwater chemistry and soils acidification processes.
9.8.2 Hydrogeology Resource Inventory
Mapping of the hydrogeological resources has been completed for the Province by
the Alberta Research Council. Groundwater chemistry varies markedly in Alberta, because
of the differences in geology, topography, climate, vegetation, and soils (Campbell
1987). The Hydrogeological Maps of Alberta contain regional information on the chemistry
of groundwater but these are often from isolated samples taken during water well
installation, logging or during soils surveys, for example. There is at present no
systematic groundwater sampling network to monitor groundwater chemistry in the Province.
9.9 SULPHUR ISOTOPE STUDIES IN ALBERTA
The following discussion was extracted from the review by Krouse (1987) on
"Sulphur isotope studies in Alberta in reference to acidic deposition".
Environmental sulphur isotope studies have been conducted in Alberta since the
late 1960's. To date, thousands of analyses have been carried out for sulphur compounds
on samples from the atmosphere, hydrosphere, pedosphere, and biosphere. From the
viewpoint of using stable isotopes to trace pollutant sulphur (S) in the environment.
Alberta is one of the few places in the world where industrial emissions differ greatly
in isotopic composition from those of the preindustrial environments. Consequently,
these investigations have not only served to identify and trace industrial S at loca-
tions in Alberta, but have contributed immensely to our understanding of fundamental
concepts concerning uptake and utilization of sulphur by environmental receptors.
The importance of sulphur isotope studies in environmental research is that
stable isotope abundances can usually provide information on the source of sulphur
pollutants. Very few other analytical techniques have this capability. Conventional
measurements of pollutant concentrations fail to apportion sources. Cases can be cited
where high concentrations of sulphate have been wrongly attributed to a specific
industrial source as a result of conventional analyses.
Basic principles of stable isotopes which pertain to sulphur pollution include
the following:
1. Isotopes of an element differ in their masses. Since many processes are
mass dependent, the relative abundances (the ratio of the number of
^""S atoms to ^^S atoms) in natural components are altered.
2. In nature the process which alters sulphur isotope abundances most sig-
nificantly is bacterial S04^~ reduction during which ^^S04^~ is converted
faster than ^'*S04^~ to sulphide.
3. Sulphur pollutants usually differ in isotopic composition from their
environmental receptors.
227
4. On the basis of (3), the presence of pollutant sulphur in the environment,
and often the ratio of pollutant-to-natural sulphur, may be determined.
5. Successful isotope tracing of pollutant sulphur requires minimal isotope
fractionation during transport and deposition. This criterion is diffi-
cult to meet in anaerobic environments because of bacterial reduction
processes. However, most studies of the atmosphere, water, and soil
involve aerobic conditions where the following processes have minimal
isotopic selectivity:
a. chemical or bacterial oxidation. Addition of oxygen does not greatly
influence the sulphur isotope composition.
b. high temperature processes, e.g., in power plant stacks. Isotope
fractionation decreases with increasing temperature.
c. S04^ assimilation by bacteria or plants.
d. reactions of sulphur compounds in the solid state, such as dissolu-
tion of evaporites or oxidation of elemental sulphur. The reaction
proceeds essentially layer by layer, thus limiting isotopic
selectivity.
e. conversions among larger complex molecules if bond rupture involves
large fragments of the molecule. In that case, isotopic substitution
does not produce a large percent change in mass in the species
undergoing reaction.
6. An exception to the above principles is the isotopically selective emis-
sion of reduced sulphur compounds by vegetation under stress. Therefore,
enrichment of heavier sulphur isotopes in vegetation to levels above those
of known sources can serve as a stress indicator.
7. Since the uniform isotope composition of O2 in the atmosphere differs
greatly from the variable isotopic composition of water, oxygen isotope
measurements of sulphate provide information concerning the oxidation of
pollutant sulphur.
8. Isotopic data should be considered as a complementary rather than an
alternate tool. With the exception of the emission of reduced sulphur
compounds by stressed vegetation, isotope data alone do not relate
environmental impacts to sources. They must be used in combination with
biological data.
9. Isotope data are most effective when background measurements are taken
prior to commencement of an industrial operation. This is seldom the case.
However, it is often possible to estimate the background conditions.
Further, it is useful at any time to establish an isotopic "baseline" with
which future measurements might be compared.
Further isotopic data may identify an important phenomenon which is not discernible from
conventional concentration data. For example, unusual enrichments of the heavier
^*S in foliar S are consistent with isotopically selective emissions of reduced S
under environmental stress. The concentrations of S alone in foliage would not provide
evidence of this biochemical reaction.
228
Studies in Alberta using isotopes as tracers have significantly advanced global
investigations of anthropogenic S in the environment. These studies have been summarized
in Fritz and Pontes (1980) and Tabatabai (1986). A review of these studies was long
overdue and, therefore, much of the review by Krouse (1987) contains data from many
studies that have not previously been published. A summary of the principal findings
gleaned from many of the Alberta studies is provided below:
1. Hydrogen sulphide released to the atmosphere from springs in the Rocky
Mountains of Alberta is depleted in the heavier sulphur isotopes as the
consequence of bacterial SOa^ reduction (Krouse et al. 1970).
2. The first study of sulphur isotope abundances in soils revealed 5^^S
values near 0°/oo in central Alberta and as low as -30°/oo in the Peace
River area (Lowe et al. 1971). 6^"$ values near 0°/oo have since been found
for soils in many locations, e.g., Ram River (Krouse 1977b) and West
Whitecourt study area (Krouse et al. 1984). Very negative values have
been found at Teepee Creek (Krouse and Case 1981) and near small lakes in
the Twin Butte area.
3. Dissolved sulphate in the Mackenzie River system was found to vary in
sulphur isotope composition over almost the total range encountered
globally for fresh water. Rivers draining the Peace River area have
S04^ with highly negative 6^*S values consistent with soils data for that
region (Hitchon and Krouse 1972).
4. In contrast to the above, sour gas (HsS-rich) in carbonate reservoirs of
Alberta has quite positive 6^"$ values (Krouse 1977a).
5. The air in Alberta near sour gas plants tends to have SO2 enriched in
^''s (i.e., positive 5^*S values) in comparison to the global average.
6. Ground level SO2 near Crossfield, Alberta became more enriched in ^'*S as
the sour gas processing plant went from shutdown to full production. The
6^*5 value (+29 °/oo) of the stack gas was mathematically predicted from
the ground level data and experimentally verified using a helicopter
mounted high volume sampler (Krouse 1980).
7. Effects of wind direction on the isotopic composition of SO2 reaching a
monitoring site were documented in the West Whitecourt Case Study (Krouse
et al. 1984). A wind-direction activated array of high volume samplers
was used.
8. The presence of sulphur of industrial origin was documented isotopically
in surface waters near sour gas processing plants, e.g., Valleyview (Krouse
and Case 1983) and the West Whitecourt study area (Krouse et al. 1984).
9. A correlation was found between the sulphur isotope composition of
S04^ and the dissolved organic S content in surface waters in the
West Whitecourt case study, indicating that industrial S interacted with
organic matter in the environment (Krouse et al. 1984).
see Krouse (1987), p. 8, Figure 3
229
10. In the Calgary river system, lateral mixing of S04^ is slow. Dif-
ferent sources of effluents identified by a cross-sectional isotopic study
is a possibility that has been suggested by Krouse (1980).
11. Random sampling in the Ram River area over two years showed that epiphytic
lichens had a 5^"$ distribution similar to atmospheric SO2 (Krouse
1977b) .
12. Lichens in the Fox Creek area had higher 5^*5 values in locations
more exposed to SO2 emissions; those on the lee side of hills had
lower values than those exposed on the windward side (Case and Krouse
1980) .
13. In contrast to statement 11, conifer needles had 6^*5 values inter-
mediate to those of atmospheric SO2 and the soil. This demonstrated
that foliage could acquire sulphur from the atmosphere as well as by
translocation from the root system (Krouse 1977b). This phenomenon has
been verified by several other investigators (Krouse 1987).
14. Foliar uptake of sulphur from the atmosphere and soil was demonstrated
isotopically in the field with potted plant experiments conducted by
Winner et al . (1978) .
15. The sulphur isotopic compositions of mosses were found to vary with
downwind distance and direction near a sour gas plant in the Fox Creek
area. Higher 5^*S values corresponded to closeness to the source of
emissions (Winner et al . 1978).
16. Ground lichens and mosses appeared not to be as enriched in as
those on trees in the same area, e.g., Brazeau (Latonas et al. 1986).
17. Moss was found capable of trapping atmospheric S compounds, thus prevent-
ing their transport to the subsoil (Krouse 1980). Litter generally is
capable of blocking the downward movement of wet and dry deposition of
S-compounds (Legge et al . 1986).
18. Conifer needles may display more positive i^'^S values with increasing
height on a given tree (unpublished data, Ram River 1972; Krouse et al.
1984). This can be explained because upper branches exert a canopy
resistance resulting in the exposure of lower branches to reduces S.
concentrations (Lester et al . 1986).
19. Consistent with statement 18, the needles on the uppermost branches of
lodgepole pine were found to be unusually enriched in ^'^S compared
with those lower in the canopy (Krouse et al. 1984).
20. Some lichens in the Fox Creek area had unusually high 5^*S values,
suggesting that under sulphur stress, gaseous compounds depleted in
5^'*S were emitted by the lichens (Case and Krouse 1980).
21 . Laboratory experiments showed that H2S was emitted by cucumber plants
grown in high concentrations of SOa^ and HS03~. The H2S was depleted in
^"s by as much as 15°/oo, compared with the nutrient solutions (Winner et
al. 1981).
230
22. In the leepee Creek area of Alberta, sulphur isotope data revealed that
foliage and soils with high sulphur contents were associated with shallow
subsurface sulphate mineral deposits. These natural sulphur sources,
quite depleted in ^"s, dominated the environmental sulphur cycle in
that area (Krouse and Case 1981).
23. Sulphur isotope analyses of soil cores near Valleyview, Alberta were used
to document penetration of sulphur of industrial origin into the subsoil
in the vicinity of a flare stack which had been operational for over two
decades (Krouse and Case 1983). Data for the West Whitecourt study area
showed subsoil movement of sulphur of industrial origin (Legge et al.
1986) .
24. At leepee Creek, soil texture was found to be an important factor with
clay particles retaining S -compounds to a greater extent than sand (Krouse
and Case 1983) .
25. Other factors influencing the penetration of industrial S into the subsoil
are vegetation cover (Statement 17), duration of emissions, and hydrology
(Krouse et al . 1984) .
26. In soil profile studies, isotope data revealed that sampling by horizon is
more meaningful than sampling pre-selected depth intervals.
21. Sulphur isotope data from the Zama area and elsewhere strongly suggest
that in rolling terrain, sulphate minerals accumulating in depressions may
contribute to visible stress symptoms on vegetation, whereas plants on
knolls may be growing in S-deficient soil. The latter may display the
isotopic signature of industrial emissions and may actually be utilizing
available atmospheric sulphur (Krouse and Case 1982).
28. As sulphur is passed through the food chain, the isotopic discrimination
is minimal, i.e., animals have a sulphur isotope composition similar to
their diets. Citizens of Calgary were found to have remarkably
consistent A^^S values (near 0°/oo) in their hair, nails,
blood, kidney stones, and urine.
231
9.10 ACIDIC DEPOSITION IN THE ALBERTA CONTEXT: LIIERATURE CIIED
Addison, P. A. 1984. Quantification of branch dwelling lichens for the detection of air
pollution impact. Li chenol ogi st 16(3): 297-304.
Addison, P. A. and K.J. Puckett. 1980. Deposition of atmospheric pollutants as measured
by lichen element content in the Athabasca Oil Sands area. Canadian Journal of
Botany 58: 2323-2334.
Addison, P. A., S.S. Malhotra, and A. A. Khan. 1984. Effect of sulfur dioxide on woody
boreal forest species grown on native soils and tailings. Journal of Environ-
mental Quality 13(3): 333-336.
Alberta. Agriculture. 1982. 1981 Census of Agriculture for Alberta. Edmonton, Alberta.
217 pp.
Amundson, R.G., R.B. Walker, and A.H. Legge.
forest: The West Whitecourt case
Air, and Soil Pollution 29: 129-147.
1986. Sulphur gas emissions in the boreal
study. VII. Pine tree physiology. Water,
Baker, J. 1977. Nutrient levels in rainfall, lodgepole pine foliage and soils surrounding
two sulfur gas extraction plants in Strachan, Alberta. Canadian Forest Service,
Information Report NOR-X-194. Edmonton, Alberta: Northern Forest Research
Centre. 18 pp.
Baker, J., D. Hocking, and M. Nyborg. 1977. Acidity of open and intercepted precipitation
in forests and effects on forest soils in Alberta, Canada. Water, Air, and
Soil Pollution 7: 449-460.
Baker, J., D. Hocking, and M. Nyborg. 1976. Acidity of open and intercepted precipitation
in forests and effects on forest soils in Alberta, Canada. I_n: Proceedings of
the First International Symposium on Acid Precipitation and the Forest Ecosys-
tem. US Department of Agriculture, Forest Service General Technical Report
NE-23, eds. L.S. Dochinger and T.A. Seliga. 1 975 May 12-15; Columbus, Ohio;
pp. 779-790.
Bloom, P.R. and D.F. Grigal. 1985. Modeling soil response to acidic deposition in non-
sulfate adsorbing soils. Journal of Environmental Quality 14: 489-495.
Bloomfield, C. 1967. Effect of some phosphate fertilizers on the oxidation of elemental
sulfur in soil. Soil Science 115: 219-223.
Bryant, R.D., E.J. Laishley, C.L. Labine, and J.B. Hyne. 1985. Thibacilli and wet sul-
phur corrosion. ASR Quarterly Bulletin XXII: 1-11.
Bryant, R.D., K.M. McGroarty, J.W. Costerton, and E.J. Laishley. 1983. Isolation and
characterization of a new acidophilic Thi obaci 1 1 us species (T. albertis) .
Canadian Journal of Microbiology 29: 1159-1170.
Campbell, K.W. 1987. Pollutant Exposure and Response Relationships: A Literature Review.
Prep for the Acid Deposition Research Program by Subsurface Technologies and
Instrumentation Limited, Calgary, Alberta, Canada. ADRP-B-07-87 . 151 pp +
maps .
Case, J.W. and H.R. Krouse. 1980. Variations in sulphur content and stable sulphur
isotope composition of vegetation near a SO2 source at Fox Creek, Alberta,
Canada. Oecologia 44: 248-257.
Costerton, J.W. and R.T. Irvin. 1981. The bacteria! glycocalyx in nature and disease.
Annual Review of Microbiology 351: 299-324.
Environment Canada. 1986. Emissions Inventory - 1986. Inventory Management Division.
Draft Report.
Environment Canada. Water Quality Branch, Data and Instrumentation Section. 1982.
NAQUADAT. Dictionary of Parameter Codes.
232
Erickson, P.K., and D.O. Trew. 1987. An assessment of the potential sensitivity of
Alberta lakes to acid deposition. Ijl' Acid Forming Emissions in Alberta and
their Ecological Effects. 2nd Symposium Workshop Proceedings, eds. H.S.
Sandhu, A.H. Legge, J.I. Pringle, and S. Vance. 1986 May 12-15; Calgary,
Alberta; 309-335.
Fritz, P. and J. Ch. Fontes, eds. 1980. Handbook of Environmental Isotope Geochemistry,
Volume 1. The Terrestrial Environment, A. New York: Elsevier Scientific
Publishing. 545 pp.
Green, R. 1982. Geological Map of Alberta. Alberta Research Council. Map 35.
Guderian, R. and H. Strattmann. 1968. Frei landversuche zur Ermittlung von Schwef eldiox-
idwirkungen auf die Vegetation. III. Teil Grenzwerte schadlicher SO2-
Immissionen fur Obstund Forstkul turen sowie fur landwi rtschaf tl iche und
gartnerische Pf lanzenarten . Forschungsberi chte des Landes Nordrhei nWestf a1 en .
114 pp.
Harvey, G.W. and A.H. Legge. 1979. The effect of sulfur dioxide upon the metabolic level
of adenosine triphosphate. Canadian Journal of Botany 57(7) :759-764.
Hesslein, R.H. 1979. Lake acidification potential in the Alberta Oil Sands. Environmen-
tal Research Program Study Area. Prepared for the Alberta Oil Sands Environ-
mental Research Program by the Freshwater Institute, Environment Canada.
AOSERP Report 71 . 36 pp.
Hill, A.C., H.E. Heggestad, and S.L. Linzon. 1970. Ozone. Iin: Recognition of Air Pol-
lution Injury to Vegetation: a Pictorial Atlas, eds. J.S. Jacobson and A.C.
Hill. Pittsburgh, Pennsylvania: Air Pollution Control Association, pp. B1-B22.
Hitchon, B. and H.R. Krouse. 1972. Hydrogeochemi stry of the surface waters of the
Mackenzie River drainage system, Canada: III. Stable isotopes of oxygen,
carbon, and sulphur. Geochimica et Cosmochimica Acta 36: 1337-1357.
Holowaychuk, N. and R.J. Fessenden. 1987. Soil sensitivity to acid deposition and the
potential of soil and geology in Alberta to reduce acidity of acidic inputs.
Earth Science Report 87-1. Edmonton: Alberta Research Council. 38 pp. 2 maps.
Holowaychuk, N. and J.D. Lindsay. 1982. Distribution and relative sensitivity to acidi-
fication of soils. Sand River area. Alberta. Prep, for the Research Management
Division, Alberta Environment and Canadian Petroleum Association. RMD Report
82/13. Edmonton, Alberta. 93 pp.
Ivarson, K.C. 1977. Changes in decomposition rate, microbial population and carbohydrate
content of an acid peat bog after liming and reclamation. Canadian Journal of
Soil Science 7: 1 29 -1 37.
Jaques, D.R. 1987. Major Biophysical Components of Alberta. Prep for the Acid Deposi-
tion Research Program by Ecosat Geobotanical Surveys Inc. ADRP-B-06/B7 .
108 pp.
Krouse, H.R. 1987. Environmental Sulphur Isotope Studies in Alberta: A Review. Prep
for the Acid Deposition Research Program by The Department of Physics, The
University of Calgary, Alberta, Canada. ADRP-B-03-87 . 89 pp.
Krouse, H.R. 1977a. Sulphur isotope studies and their role in petroleum exploration.
Iin: Application of Geochemistry to the Search for Crude Oil and Natural Gas,
ed . B. Hitchon. Special Issue, Journal of Geochemical Exploration I. New
York: Elsevier Scientific Publishing, pp. 189-211.
Krouse, H.R. 1977b. Sulphur isotope abundance elucidate uptake of atmospheric sulphur
emissions by vegetation. Nature 265: 45-46.
Krouse, H.R. 1980. Sulphur isotopes in our environment. In: Handbook of Environmental
Isotope Geochemistry, eds. P. Fritz and J.Ch. Fontes. New York: Elsevier
Scientific Publishing, pp. 435-471.
233
Krouse, H.R. and J.W. Case. 1983. Sulphur isotope abundances in the environment and
their relation to long term sour gas flaring, near Valleyview, Alberta.
Prepared for Alberta Environment, Research Management Division and Pollution
Control Division, Air Quality Control Branch by Department of Physics,
University of Calgary. RMD Report 83/18. 110 pp.
Krouse, H.R. and J.W. Case. 1982. Sulphur isotope tracing of acid emissions into the
environment in Alberta. In: Proceedings of the Symposium on Acid Forming
Emissions in Alberta and Their Ecological Effects. 1982 March 9-12; Edmonton,
Alberta. Alberta Department of the Environment, Canadian Petroleum
Association, Oil Sands Environmental Study Group; pp. 207-232.
Krouse, H.R. and J.W. Case. 1981. Sulphur isotope ratios in water, air, soil and vege-
tation near Teepee Creek gas plant. Alberta. Water, Air, and Soil Pollution
15: 11 .
Krouse, H.R., A.H. Legge, and H.M. Brown. 1984. Sulphur gas emissions in the boreal
forest: The West Whitecourt Case Study V: Stable Sulphur Isotopes. Water,
Air, and Soil Pollution 22: 321-347.
Krouse, H.R., F.D. Cook, A. Sasaki, and V. Smejkal. 1970. Microbiological isotope frac-
tionation in springs of western Canada. In^: Recent Developments in Mass
Spectroscopy, eds. K. Ogata and T. Hayakawa. Chicago, Illinois: University
Park Press, pp. 629-639.
Ladd, T.I. 1982. Heterotrophic activity of epilithic bacteria communities in a lotic
ecosystem. Ph.D. Thesis. Calgary, Alberta: University of Calgary. 402 pp.
Laishley, E.J. and R.D. Bryant. 1985. Long term environmental effects of sulphur and
sulphur concrete on the soil ecosystem. In: Proceedings of Sulphur-84, an
International Conference on Sulphur, Calgary, Alberta. 1984 June 3-6; Sulphur
Institute of Canada; pp. 641-646.
Laishley, E.J. and R. Bryant. 1987. Critical Review of Inorganic Sulphur Microbiology
with Particular Reference to Alberta Soils. Prep for the Acid Deposition
Research Program by the Department of Biology, The University of Calgary.
ADRP-B-04/87. 56 pp.
Latonas, G.P., H.R. Krouse, G. Collin, and L. Lewis. 1986. Application of sulphur isotope
analysis in environmental monitoring programs for sour gas processing
operations. A case study of the Petro-Canada Brazeau project. In: Northern
Hydrocarbon Development, Environmental Problem Solving, ed . M.L. Lewis.
Proceedings of the 8th Annual Meeting of the Society of Petroleum Industry
Biologists Conference, Banff, September 1985. University of Toronto Press,
Toronto, pp. 99-111.
Legge, A.H. 1982. Sulphur gas emissions in the boreal forest: The West Whitecourt case
study. Water, Air, and Soil Pollution 17: 379-398.
Legge, A.H. and J.C. Bogner. 1983. Ecological monitoring of sulphur in forests in west-
ern Canada. Aquilo Series Botanica 19: 119-139.
Legge, A.H., J. Corbin, J. Bogner, M. Strosher, D. Parkinson, H.R. Krouse, E.J. Laishley,
M.J. Cavey, C.E. Prescott, R.J.F. Bewley, M. Nosal, H.U. Schellhase, T.C.
Weidensaul, and J. Mayo. 1986. Acidification in a temperate forest ecosystem:
The role of sulphur gas emissions and sulphur dust. A report submitted to the
Whitecourt Environmental Study Group. 471 pp.
Legge, A.H., H.C. Kaufmann, and J.W. Winchester. 1984. Tree ring analysis by PIXE for a
historical record of soil chemistry response to acidic air pollution. Nuclear
Instruments and Methods in Physics Research B3: 507-510.
Legge, A.H., D.R. Jaques, G.W. Harvey, H.R. Krouse, H.M. Brown, E.C. Rhodes, M. Nosal,
H.U. Schellhase, J. Mayo, A. P. Hartgerink, P.F. Lester, R.G. Amundson, and R.B.
Walker. 1981. Sulphur gas emissions in the boreal forest: The West Whitecourt
case study I. Executive summary. Water, Air and Soil Pollution 15: 77-85.
234
Legge, A.H., E.M. van Zinderen Bakker, Jr., E. Peake, and D.C. Lindsay. 1980. The oxides
of nitrogen and their interactions in the environment: a review. A Final
Report Submitted to the Research Secretariat, Alberta Environment, and Canadian
Petroleum Association by Kananaskis Centre for Environmental Research,
University of Calgary. Calgary, Alberta. 169 pp.
Legge, A.H., D.R. Jaques, R.G. Amundson, and R.B. Walker. 1977. Field studies of pine,
spruce and aspen periodically subjected to sulfur gas emissions. Water, Air,
and Soil Pollution 8: 105-129.
Lester, P.F., E.C. Rhodes, and A.H. Legge. 1986. Sulphur gas emissions in the boreal
forest: the West Whitecourt case study. IV: Air quality and the
meteorological environment. Water, Air, and Soil Pollution 27: 85-108.
Lore, J. A. 1984. Observations of pH in field soils near a sour gas plant in Alberta.
Water, Air, and Soil Pollution 21: 403-410.
Lowe, L.E., A. Sasaki, and H.R. Krouse. 1971. Variations of sulphur-34: sulphur-32 ratios
in soil fractions in western Canada. Canadian Journal of Soil Sciences 51:
129-131 .
Malhotra, S.S. 1977. Effects of aqueous sulfur dioxide on chlorophyll destruction in
Pinus contorta. New Phytologist 78: 101-109.
Mayo, J.M. 1987. The Effects of Acid Deposition on Forests. Prep for the Acid Deposition
Research Program by the Department of Biology, Emporia State University.
ADRP-B-09.87. 80 pp.
National Academy of Sciences. 1977. Nitrogen oxides. Committee on Medical and Biological
Effects of Environmental Pollutants. Washington, D.C. 333 pp.
Nyborg, M. J. Crepin, D. Hocking, and J. Baker. 1977. Effect of sulphur dioxide on pre-
cipitation and on the sulphur content and acidity of soils in Alberta, Canada.
Water, Air, and Soil Pollution 7: 439-448.
Nyborg, M. and P.B. Hoyt. 1978. Effects of soil acidity and liming on mineralization of
soil nitrogen. Canadian Journal of Soil Science 58: 331-338.
Penney, D.C, M. Nyborg, P.B. Hoyt, W.A. Rice, B. Siemens, and D.H. Laverty. 1 977. An
Assessment of the soil acidity problem in Alberta and Northeastern British
Columbia. Canadian Journal of Soil Science 57: 157-164.
Pielou, E.C. 1982. Testing for vegetation gradients in the Athabasca Oil Sands area.
In : Acid Forming Emissions in Alberta and Their Ecological Effects, eds. H.S.
Sandhu, J.R. Clements, and B.L. Magill. 1982 March 9-12; Edmonton, Alberta,
pp. 435-448.
Sanderson, K. 1984. Acid-forming emissions. Transportation and effects. Environment
Council of Alberta. 53 pp.
Schwartz, F.W. 1980. Hydrogeological investigation of Muskeg River Basin, Alberta.
Alberta Department of Environment and Environment Canada, Edmonton, Alberta.
Alberta Oil Sands Environmental Research Program Project WS 2.2. 97 pp.
Schwartz, F.W. 1979. Interim report on a hydrogeological investigation of the Muskeg
River Basin, Alberta. Alberta Department of Environment and Environment
Canada, Edmonton, Alberta. Prepared for Alberta Oil Sands Environmental
Research Program. 92 pp.
Shewchuk, S.R. 1981. Regional transport of air pollutants and associated ecosystem
sensitivities within a selected area of western Canada. Saskatchewan Research
Council Publication No. P-780-6-A-81 . 33 pp.
Strong, W.L. and K.R. Leggat. 1981. Ecoregions of Alberta. Alberta Energy and Natural
Resources. Report number T/4. 64 pp.
235
Tabatabai, M.A., ed . 1986. Sulfur in Agruculture. Agronomy Series No. 27. American
Society of Agronomy, Inc., Crop Science Society of America, Inc., and Soil
Science Society of America, Inc. Madison, Wisconsin. 668 pp.
Takakawa, S., T. Fujimori, and H. Iwasaki . 1979. Some properties of cell-sulfur adhesion
in Thiobaci 1 1 us thiooxidans . Journal of General and Applied Microbiology 25:
21-29.
Telang, S.A. 1987. Surface Water Acidification Literature Review. Prep, for the Acid
Deposition Research Program by the Kananaskis Centre for Environmental
Research. The University of Calgary, Calgary, Alberta. ADRP-B-01 -87 . 123 pp.
Torn, M.S., J.E. Degrane, and J.H. Shinn. 1987. The Effects of Acidic Deposition on
Alberta Agriculture: A Review. Prepared for the Acid Deposition Research
Program by the Environmental Sciences Division, Lawrence Livermore National
Laboratory. ADRP-B-08/87 . 160 pp.
Toth, J. 1963. A theoretical analysis of groundwater flow in small drainage basins.
Journal of Geophysical Research 68: 4795-4812.
Turchenek, L.W., S.A. Abboud, C.J. Tomas, R.J. Fessenden, and N. Holowaychuk. 1987.
Effects of Acid Deposition on Soils in Alberta. Prep for the Acid deposition
Research Program by the Alberta Research Council, Edmonton. ADRP-B-05-87 .
Visser, S, R.M. Danielson, and J.F. Parr. 1987. Effects of Acid-Forming Emissions on
Soil Microorganisms and Microbial ly-Mediated Processes. Prep for the Acid
Deposition Research Program by the Kananaskis Centre for Environmental
Research, The University of Calgary, and U.S. Department of Agriculture,
Beltsville, Maryland. ADRP-B-02-87 . 86 pp.
Winner, W.E., C.L. Smith, G.W. Koch, H.A. Mooney, J.D. Bewley, and H.R. Krouse. 1981.
Rates of emission of H2S from plants and patterns of stable sulfur isotope
fractionation. Nature 289: 672-673.
Winner, W.E., J.D. Bewley, H.R. Krouse, and H.M. Brown. 1978. Stable sulphur isotope
analysis of SO2 pollution impact on vegetation. Oecologia 36: 351-361.
ISBN 0-921625-13-8 (Volume 11)
ISBN 0-921625-02-2 (Set of 11)
Live Music
Archive
Librivox
Free Audio
Metropolitan Museum
Cleveland
Museum of Art
Internet
Arcade
Console Living Room
Open Library
American
Libraries
TV News
Understanding
9/11